Biomethane potential test for rapid assessment of anaerobic ...

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Biomethane potential test for rapid assessment ofanaerobic digestion of sewage sludge: co-digestionwith glycerol and trace organic removalThanh NguyenUniversity of Wollongong

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Recommended CitationNguyen, Thanh, Biomethane potential test for rapid assessment of anaerobic digestion of sewage sludge: co-digestion with glycerol andtrace organic removal, Master of Philosophy thesis, School of Civil, Mining and Environmental Engineering - Faculty of Engineeringand Information Sciences, University of Wollongong, 2014. http://ro.uow.edu.au/theses/4398

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Faculty of Engineering and Information Sciences

School of Civil, Mining and Environmental Engineering

Biomethane potential test for rapid assessment of anaerobic

digestion of sewage sludge: co-digestion with glycerol and trace

organic removal

Thanh Nguyen

This thesis is presented as part of the requirements for the

award of the Degree of Master of Philosophy

from the University of Wollongong

March 2014

i

CERTIFICATION

I, Thanh Nguyen, hereby declare that this thesis, submitted in partial fulfilment of the

requirements for the award of Master of Engineering by Research Degree, to the

school of Civil, Mining and Environmental Engineering, Faculty of Engineering,

University of Wollongong is wholly my own work unless otherwise referred or

acknowledged. The document has not been submitted for qualification at any other

academic institution.

Thanh Nguyen

March 2014

ii

ABSTRACT

Anaerobic digestion (AD) is the most widely used treatment process for sewage

sludge stabilisation over concerns of public health. In addition, the production of

methane (CH4), a renewable fuel, has also shaped the prospective of AD within the

context of energy security and global warming. This dissertation thus aimed to

evaluate two main aspects of the AD process of sewage sludge. First, the potential of

enhancing biogas production was assessed when co-digesting sewage sludge with

glycerol. Glycerol is a by-product of biodiesel production and then rich in organic

carbon. Second, the capacity of removing trace organic compounds (TrOCs) was

examined.

Results reported here suggest the stability of anaerobic conversion determined the

ultimate CH4 yield, greatly affecting the assessment of CH4 potential. Alkalinity

buffer and inoculum over substrate (I/S) ratio were demonstrated to be important

factors in maintaining steady state of the AD process. The addition of NaHCO3

resulted in an increase of CH4 production of sewage sludge that could be ascribed to

well-buffered conditions for methanogens. A significant enhancement in CH4 yield

from sewage sludge was achieved with an increase of I/S ratio from 1/9 to 1/1.

Adequate quantity of inoculum for degrading substrate was responsible for such

improvement. Moreover, no disruption of CH4 production was found since only

0.25% and 0.5% glycerol of digested sludge (by volume) was applied. The obtained

ultimate CH4 yields of three types of glycerol were comparable, indicating that any

glycerol was possibly used as a potential co-substrate of sewage sludge.

Additionally, the identification of appropriate values of alkalinity and I/S ratio for an

optimal CH4 production was dependent on key characteristics (i.e. pH and alkalinity)

of sewage sludge and glycerol.

Experimental results from anaerobic co-digestion of raw primary sludge and glycerol

show that the addition of 0.5% and 1.0% of glycerol to raw primary sludge (by

volume) could improve the anaerobic sludge conversion in terms of the daily and

cumulative CH4 production. Carbon-rich content and high solubility of glycerol led

to an increase only within the first seven days. Despite the buffer supplementation,

the instability of the AD process was clearly observed in this batch test through

characteristics of the feeds and digestates (acidic pH and low alkalinity for example),

iii

suggesting a shock of organic loading due to insufficient inoculum and extra organic

matter from glycerol. The feasibility of glycerol as a co-substrate of raw primary

sludge was satisfactory to give rise to CH4 generation. On the other hand, excessive

addition of glycerol as a co-substrate may destabilise the AD process, and thus be

counter-productive.

The removal of TrOCs by AD was evaluated. The evolution of pH and alkalinity as

well as CH4 yields obtained from experiment data and kinetic analysis indicated a

steady state of methanogenesis in TrOC-spiked bottles after 1.6 days. In a two-phase

matrix (i.e. sludge and water), TrOCs could partition (adsorb) to the solid phase as

well as the water phase. Their distribution in these phases was examined using a two-

phase fate model, which considers mass transfer (i.e. sorption and desorption). In this

study, the removal of the investigated TrOCs is defined as anaerobic biodegradation.

Their various removal efficiencies were consistently related to their molecular

properties in terms of functional groups rather than other physicochemical properties.

Compounds with the inclusion of halogen, methoxy and amine/amide groups

exhibited an effective removal while low or no removal efficiencies were observed in

compounds possessing alkyl functional groups. Regarding compounds possessing

high biodegradability (kb> 0.001 L/gVS.d), water-solid partition coefficient (kp)

determined from the proposed two-phase fate model could describe their distribution

in the sludge phase better than their hydrophobicity (in terms of log D). For

biologically persistent compounds (kb< 0.001L/gVS.d), sorption properties, which

can be predicted by log D values, were responsible for the fate of organic

contaminants in sludge matrix.

An economic design at laboratory scale of biomethane potential testing could rapidly

assess the AD of sewage sludge. Results also demonstrated increased methane yields

when using glycerol as a co-substrate of sewage sludge and trace organic removal

efficiency under anaerobic conditions. These findings are of great importance for

scaling up AD facilities in real wastewater treatment plants.

Keywords: anaerobic digestion, anaerobic co-digestion, biomethane potential,

sewage sludge, glycerol, trace organic compounds, sorption.

iv

ACKNOWLEDGEMENTS

Completion of my Master by Research at the University of Wollongong has not been

achieved by my efforts alone, but memorably contributed by many wonderful people

to whom I must express my honest thanks.

My sincere gratitude is offered to Professor Long Duc Nghiem who gave me a

precious opportunity to carry out my research in his Strategic Water Infrastructure

Laboratory along with his enthusiastic guidance and support, which led to significant

improvement in my academic writing and research skills. I must also sincerely thank

Associate Professor Muttucumaru Sivakumar, my co-supervisor for his insightful

comments and suggestions throughout my course.

A word of thanks must be also recorded to other team members associated with the

Strategic Water Infrastructure Laboratory for their commitment and companionship

throughout my study. I would also like to offer special thanks to Dr. Jinguo Kang for

his help regarding trace organic analyses and gratitude to Kaushalya Wijekoon for

her suggestions and assistance in some of the laboratory analyses and system

operations.

I would also express my thanks to our laboratory technicians, Mr. Robert Rowlan,

Mr. Frank Crabtree and our laboratory manager Dr. Ling Tie for their whole-hearted

assistance from the very first days of setting up my experiment system.

Thanh Hoa Province Government, Vietnam and the University of Wollongong are

greatly thanked for offering me the Master scholarship.

To my family, my heartfelt thanks are expressed for their unconditional love and

belief.

v

TABLE OF CONTENTS

CERTIFICATION ........................................................................................................ i

ABSTRACT ................................................................................................................. ii

ACKNOWLEDGEMENTS ........................................................................................ iv

TABLE OF CONTENTS ............................................................................................. v

LIST OF FIGURES .................................................................................................. viii

LIST OF TABLES ..................................................................................................... xii

LIST OF ABBREVIATION ..................................................................................... xiv

1 INTRODUCTION .................................................................................................... 1

1.1 Background of the study ................................................................................. 1

1.2 Scope of study ................................................................................................. 3

1.3 Research objectives ......................................................................................... 3

1.4 Expected outcomes.......................................................................................... 4

1.5 Thesis outline .................................................................................................. 4

2 LITERATURE REVIEW.......................................................................................... 6

2.1 Sewage sludge ................................................................................................. 6

2.1.1 Types of sewage sludge ........................................................................... 6

2.1.2 Constituents of sewage sludge ................................................................. 9

2.2 Anaerobic sludge digestion ........................................................................... 11

2.2.1 Fundamentals of anaerobic digestion ..................................................... 11

2.2.2 Current status of the anaerobic digestion process .................................. 11

2.2.3 Factors influencing the anaerobic digestion process .............................. 14

2.2.4 Anaerobic digestion in sewage sludge treatment ................................... 18

2.3 Anaerobic co-digestion of sewage sludge with other substrates ................... 20

2.3.1 Advantages of anaerobic co-digestion in treating sewage sludge .......... 20

2.3.2 Current status of anaerobic co-digestion of sewage sludge and other

organic materials ................................................................................................ 21

2.4 The behaviour of trace organic contaminants during the anaerobic digestion

process .................................................................................................................... 29

2.4.1 Trace organic compounds and their effects ........................................... 29

2.4.2 Occurrence of emerging contaminants in sludge ................................... 30

2.4.3 Anaerobic digestion in the removal of trace organic compounds .......... 31

2.5 Summary ....................................................................................................... 34

vi

3 MATERIALS AND METHODS ............................................................................ 35

3.1 Materials ........................................................................................................ 35

3.1.1 Inoculum ................................................................................................ 35

3.1.2 Substrates ............................................................................................... 35

3.2 Experimental methods ................................................................................... 35

3.2.1 Biomethane potential test equipment ..................................................... 35

3.2.2 Experimental protocols .......................................................................... 36

3.2.3 Experimental description ....................................................................... 38

3.3 Analytical methods........................................................................................ 42

3.3.1 Basic parameters .................................................................................... 42

3.3.2 Chemical oxygen demand ...................................................................... 43

3.3.3 Methane yield calculation ...................................................................... 43

3.3.4 Trace organic analysis ............................................................................ 45

3.3.5 Viscosity of glycerol .............................................................................. 46

4 BIOMETHANE POTENTIALOF SEWAGE SLUDGE AND GLYCEROL ........ 48

4.1 Sewage sludge ............................................................................................... 48

4.1.1 Characteristics of sewage sludge ........................................................... 48

4.1.2 Effects of buffer supplement on the anaerobic treatment of sewage

sludge ................................................................................................................ 50

4.1.3 Methane production of sewage sludge ................................................... 56

4.2 Glycerol as a co-substrate ............................................................................. 62

4.2.1 Characteristics of pure and crude glycerol ............................................. 62

4.2.2 Variations of control parameters in the study of methanogenesis of

glycerol ............................................................................................................... 64

4.2.3 Methanogenic conversion of glycerol .................................................... 68

4.3 Summary ....................................................................................................... 74

5 ANAEROBIC CO-DIGESTION OF SEWAGE SLUDGE AND GLYCEROL .... 75

5.1 Characteristics of mixtures of sewage sludge and glycerol .......................... 75

5.2 Removal of volatile solids and chemical oxygen demand ............................ 79

5.3 Methane potential of co-digestion mixture of sewage sludge and glycerol .. 82

5.4 Summary ....................................................................................................... 90

6 FATE OF TRACE ORGANIC COMPOUNDS DURING ANAEROBIC

DIGESTION OF SEWAGE SLUDGE ...................................................................... 91

vii

6.1 Overall performance of AD of trace organics ............................................... 91

6.1.1 The variation of control parameters during over the digestion time ...... 91

6.1.2 Anaerobic conversion of trace organic compounds ............................... 94

6.2 Dynamics of trace organics under anaerobic sludge treatment ..................... 97

6.3 Removal performance of trace organic compounds under anaerobic sludge

treatment ............................................................................................................... 108

6.3.1 Overall removal of trace organic compounds ...................................... 108

6.3.2 Role of chemical structures .................................................................. 111

6.4 Fate of trace organic compounds in sludge matrix under anaerobic condition.

..................................................................................................................... 115

6.5 Summary ..................................................................................................... 120

7 CONCLUSIONS AND RECOMMENDATIONS ............................................... 121

7.1 Conclusions ................................................................................................. 121

7.2 Recommendations for future studies ........................................................... 123

viii

LIST OF FIGURES

Figure 1.The descriptive structure of thesis. ............................................................... 5

Figure 2. Different types of sewage sludge (modified from [10, 33]). ....................... 7

Figure 3. Different stages of AD process (modified from [4, 41, 42]). .................... 12

Figure 4. Factors influencing AD performance. ....................................................... 14

Figure 5. A picture of the BMP test system. ............................................................. 37

Figure 6. The schematic diagram of BMP test system. ............................................. 37

Figure 7. A picture of one BMP bottle unit used in this study. ................................. 38

Figure 8. Experimental roadmap. .............................................................................. 39

Figure 9. Hach equipment for COD determination: a) DBR200 COD Reactor and

b)DR/2000 Spectrophotometer. ......................................................................... 43

Figure 10. A schematic of sample preparation in solid and liquid phases GC-MS

analysis. .............................................................................................................. 46

Figure 11. A photograph of the Anton Paar Physica MCR 301 Rheometer. ............ 47

Figure 12. Cumulative CH4 yield and CH4 production rate for buffered and non-

buffered reactors with increasing buffer concentrations, 0, 15, and 30 mM

NaHCO3, respectively labelled as R-0, R-15, and R-30. .................................... 53

Figure 13. Temporal variations of cumulative CH4 production of the sample mixture

R1 (1/9 I/S) supplemented with increasing buffer concentrations, 0, 15, and 30

mM NaHCO3, respectively labelled as R-0, R-15, and R-30. ............................ 54

Figure 14. Plots of cumulative CH4 yields on COD basis and regression fitting

curves of the first order and modified Gompertz models. ................................. 55

Figure 15. Profile of methanogenesis of raw primary sludge in terms of production

and yield over experimental time; error bars refer to the standard deviation (n =

2). ....................................................................................................................... 58

Figure 16. Plots of experimental cumulative CH4 yield on VS basis and regression

fitting curves of the first order model and modified Gompertz model applied for

raw primary sludge. ............................................................................................ 60


fitting curve of the modified Gompertz model 2 applied for raw primary sludge.

............................................................................................................................ 60

Figure 18. Viscosity of pure glycerol and two types of crude glycerol (biodiesel and

BIA) as a function of temperature from 10 °C to 60 °C. ................................... 63

ix

Figure 19. Variations of total alkalinity and ratios of total organic acids and total

alkalinity as related to type and concentrations of glycerol; 0.25% and 0.5%

addition of pure (P), biodiesel (BID) and BIA glycerol were labelled as P-0.25,

P-0.5, BID-0.25, BID-0.5, BIA-0.25, and BIA-0.5 respectively; error bars refer

to the standard deviation (n=2). ......................................................................... 67

Figure 20. Cumulative CH4 yield and CH4 production rate of three different types of

glycerol, namely pure (P), biodiesel (BID) and BIA, at increasing

concentrations of 0.25% and 0.5%, respectively labelled as P-0.25, P-0.5, BID-

0.25, BID-0.5, BIA-0.25, and BIA-0.5. ............................................................. 69


curves of the first order and modified Gompertz models of pure glycerol at two

concentrations of 0.25% (P-0.25) and 0.5% (P-0.5). ......................................... 71


curves using the first order and modified Gompertz models of a) biodiesel

(BID) and b) BIA glycerol at 0.25% and 0.5% concentrations, respectively

labelled as BID-0.25, BID-0.5, BIA-0.25 and BIA-0.5. .................................... 72

Figure 23. Values of pH and alkalinity as related to the glycerol and buffer addition

before and after experimental time; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0

refer to the mixture of 10% digested sludge and 90% raw primary sludge by

volume (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and

introduced with 0.5% and 1.0% glycerol each, respectively. ............................ 77

Figure 24. Changes of VS, CODs, and ratios of VS/TS and CODs/CODt in relation

with glycerol and buffer addition; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer

to the mixture of 10% digested sludge and 90% raw primary sludge by volume

(R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and


Figure 25. Profile of solid content, including final VS concentrations, VS/TS ratio,

VS removal efficiency at the end of digestion; R-15-0.5, R-15-1.0, R-30-0.5, R-

30-1.0 refer to the mixture of 10% digested sludge and 90% raw primary sludge

by volume (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and


Figure 26. CODs concentrations and CODs/CODt ratios in digestates; R-15-0.5, R-

15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 10% digested sludge and 90%

x

raw primary sludge by volume (R-0) supplemented with 15 (R-15) and 30 mM

NaHCO3 (R-30), and introduced with 0.5% and 1.0% glycerol each,

respectively. ....................................................................................................... 81

Figure 27. Profile of daily and cumulative CH4 production in the co-digestion batch

test; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 10%

digested sludge and 90% raw primary sludge by volume (R-0) supplemented

with 15 (R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5% and

1.0% glycerol each, respectively. ....................................................................... 83

Figure 28. Profile of cumulative CH4 production on basis of VS removal

(mLCH4/gVSremoval) and COD added (mLCH4/gCODadded); R-15-0.5, R-15-1.0,

R-30-0.5, R-30-1.0 refer to the mixture of digested sludge and raw primary

sludge (1/9 by volume) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-

30), and introduced with 0.5% and 1.0% glycerol each, respectively. .............. 85


curves of the first order and modified Gompertz models; R-15-0.5, R-15-1.0

refer to the mixture of 1/9 I/S supplemented with 15 mM NaHCO3 (R-15) and

introduced with 0.5% and 1.0% glycerol respectively. ...................................... 86


curves of the first order and modified Gompertz models; R-30-0.5, R-30-1.0

refer to the mixture of digested sludge and raw primary sludge (1/9 by volume)

supplemented with 30 mM NaHCO3 (R-30) and introduced with 0.5% and 1.0%

glycerol respectively. ......................................................................................... 87

Figure 31. Profile of SMA of reference bottles and co-digestion bottles; R-15-0.5, R-

15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 1/9 I/S supplemented with 15

(R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5% and 1.0%

glycerol each, respectively. ................................................................................ 89

Figure 32. Profile of pH and alkalinity values in digested sludge bottles and spiked

digested sludge bottles at the particular digestion time. .................................... 93

Figure 33. Profile of CODs and ratio of CODs/CODt of digested sludge and digested

sludge spiked with TrOCs from ay 0 to day 35 of digestion. ............................ 94

Figure 34. Profile of daily and cumulative CH4 production of reference and TrOC-

spiked MBP bottles over 35 days; error bars refer to the standard deviation (n =

2). ....................................................................................................................... 95

xi

Figure 35. Plots of cumulative CH4 yields on VS basis and regression fitting curve

using the first order and modified Gompertz models applied for digested sludge,

TrOC-spiked digested sludge, and methanol. .................................................... 96

Figure 36. Distribution of the selected TrOCs in the water and solid phases of

digested sludge. .................................................................................................. 98

Figure 37. The schematic diagram of the fate of TrOCs in a sludge-water system. . 99

Figure 38. Plots of experimental data and fitting curvesof the two-phase fate model

for the selected TrOCs. .................................................................................... 103

Figure 39. Correlation between log D (pH = 8) and log kp of the selected TrOCs. 117

Figure 40. Relation among biodegradation rate constant (kb, L/gVS.d) water-solid

partition (kp, L/g) and log D (pH =8) of the selected TrOCs. .......................... 118

Figure 41. Correlation of log D (pH = 8) and kp of compounds showing low (a) and

high (b) biodegradability rates (kb). ................................................................. 119

xii

LIST OF TABLES

Table 1. Typical constituents of different types of sludge [2, 10]. ............................ 10

Table 2. Operational data of different anaerobic bioreactors used for different types

of agricultural waste [5, 44, 47, 50]. .................................................................. 15

Table 3. Advantages and disadvantages of anaerobic sludge digestion [5, 6, 10, 52,

62]. ..................................................................................................................... 19

Table 4. Biogas composition [9]. .............................................................................. 23

Table 5. Characteristics of certain co-substrates used for anaerobic digesting with

sewage sludge..................................................................................................... 25

Table 6. Co-digestion of sewage sludge with different substrates. ........................... 27

Table 7. Proposed limit values of TrOCs in sewage sludge [112-114]. .................... 31

Table 8. Experimental description of BMP bottles with increasing buffering

capacity. ............................................................................................................. 39

Table 9. Experimental description of BMP bottles with glycerol addition. .............. 40

Table 10. BMP bottles assessing BMP of different types of glycerol. ..................... 41

Table 11. Characteristics of raw, digested, and exhausted digested sludge. ............. 49

Table 12. Characteristics of sludge samples. ............................................................ 50

Table 13. Kinetic analysis of CH4 production in the batch test of buffer

enhancement. ...................................................................................................... 56

Table 14. Characteristics of raw primary sludge and sludge mixture. ...................... 57

Table 15. Kinetic analysis of CH4 production potential of raw primary sludge. ...... 61

Table 16. Physicochemical properties of pure and crude glycerol. ........................... 62

Table 17. Key properties of sample mixtures of three types of glycerol and sludge. 65

Table 18. Kinetic analysis of CH4 production during the batch test of BMP of

glycerol. .............................................................................................................. 73

Table 19. Physical and biochemical characteristics of sample mixtures in the

anaerobic co-digestion test. ................................................................................ 76

Table 20. Kinetic analysis of CH4 production on COD basis in the batch test of co-

digestion of raw primary sludge and glycerol. ................................................... 88

Table 21. Characteristics of digested sludge bottles and TrOC-spiked bottles

throughout the experimental time. ..................................................................... 92

xiii

Table 22. Kinetic analysis of CH4 production on VS basis in the batch test of the AD

treating trace organic contaminants. .................................................................. 97

Table 23. Calculated model parameters achieved from the two-phase fate model. 102

Table 24. Overall removal efficiencies of the selected TrOCs in comparison with

reported data under anaerobic and aerobic conditions. .................................... 109

Table 25. Relation of functional moieties and the removal efficiency of the selected

TrOCs. .............................................................................................................. 113

Table 26.Selected TrOCs and their relevant physicochemical properties. .............. 135

xiv

LIST OF ABBREVIATION

AD Anaerobic digestion

BMP Biomethane potential

COD Chemical oxygen demand

CODs Chemical oxygen demand – soluble fraction

CODt Chemical oxygen demand – total sample

EDCs Endocrine disrupting chemicals

GC–MS Gas chromatography – mass spectrometer

I/S Inoculum over substrate

MBR Membrane bioreactor

PPCPs Pharmaceutical and personal care products

SMA Specific methanogenic activity

SPE Solid phase extraction

TrOCs Trace organic compounds

TS Total solids

VFAs Volatile fatty acids

VS Volatile solids

WWTPs Wastewater treatment plants

1

1 INTRODUCTION

1.1 Background of the study

Since the 1990 North Sea Conference, where international agreement took place to

phase out the discharge of raw sewage sludge at sea, the treatment of sewage sludge

from wastewater treatment plants prior to environmental disposal has become a norm

in developed countries. There are several techniques for the treatment and

management of sewage sludge, including landfilling, incineration, composting, and

anaerobic digestion (AD) process [1, 2]. Among them, AD is the most commonly

used technique since biogas, which is a valuable form of renewable energy, can be

extracted from sewage sludge.

In recent years, the application of AD for sewage sludge treatment has grown rapidly

around the world. AD involves several sequential biochemical processes. Each stage

is consistently performed by a specific bacterial group in the absence of oxygen [3].

The production of biogas in AD offers several significant advantages over other

alternative technologies. These include biogas production, nutrient recovery, and

reduction of waste organic content and pathogen agents [4-6]. Recent development of

high-rate anaerobic systems has made AD an attractive technology for managing

organic wastes from agricultural production, as well as municipal and industrial

wastewater treatment [7]. AD can be applied to a range of feedstocks including solid

wastes from husbandry, food processing, and wastewater sludge [8, 9].

Sewage sludge can be described as a byproduct mixture of solids and water from

wastewater treatment [10]. By applying different treatment processes, the resulting

sewage sludge types completely differ in their characteristics. Constituents of sewage

sludge in terms of carbohydrate, lipids and protein are highly variable depending on

their origin [2]. The presence of significant concentrations of nitrogen, phosphorus

and potassium in sewage sludge make it possible for fertilising soil since these

elements are essential for plant growth.

AD instability is caused by fluctuations in organic loading rate, heterogeneity of

wastes or excessive inhibitors. Towards improving AD performance in biogas

production and accelerating the microbial activity for higher quality of biosolids,

various environmental conditions should be meticulously controlled. Additionally,

several studies have demonstrated that the hydrolysis phase is a rate-limiting step,

and directly affect the performance of AD [11-14].

2

Co-digestion of sewage sludge with higher degradable carbon source has become an

effective solution to enhance anaerobic degradation and biogas production. Anaerobic

co-digestion involves the simultaneous digestion of a homogenous mixture of two or

more substrates and has been promoted very recently in many WWTPs [15]. Not only

does this process accelerate the hydrolysis and biogas yield, it also offers many other

benefits, including dilution of potential toxic compounds, the supply of missing

nutrients, synergistic effects of microorganisms, increased load of biodegradable

organic matter, economic advantage of sharing equipment, and better biogas yield

[16, 17]. Of the organic waste materials investigated in the literature, crude glycerol

has been reported to be an ideal co-substrate for the anaerobic mineralisation of

sewage sludge [18-20]. Crude glycerol is highly soluble and rich in easily degradable

components.

Given that both biogas production and digestate quality from the anaerobic treatment

of sewage sludge were of great importance, considerations must also be given to the

fate of trace organic compounds (TrOCs) available in sewage sludge. The

accumulation of certain emerging organic compounds at trace levels onto sewage

sludge has been widely observed from WWTPs since adsorption is one of main

mechanisms responsible for their elimination during several wastewater treatment

processes [21-23]. For example, Samaras et al. [21] determined that while

biodegradation/biotransformation was responsible for trace organic removal from

wastewater the sorption tendency onto sludge particles contributed to the removal of

nonylphenol and triclosan from wastewater.

Studies reporting on the fate of these compounds during AD of sewage sludge have

aroused question over their occurrence in sewage sludge and their subsequent impact

on agricultural applications of biosolids. Unlike wastewater, sewage sludge was

regarded as a two-compartment matrix, including water and solid phases. As such,

under anaerobic sludge treatment process, every compound could undergo two main

routes, eliminated by abiotic and/or biotic removal mechanisms and/or accumulated

onto sludge particles. Several studies have indicated significant removal efficiencies

of some TrOCs including surfactants, pharmaceuticals, and personal care products by

AD treatment [22-27]. The author mainly attributed such efficient removal

performance to anaerobic biodegradation. Nonetheless, the behaviour of TrOCs in

two phases of sewage sludge and the relative impact of their behaviour on

3

biodegradability remained unclear. An extension of investigations on compounds

belonging to different usage groups with diverse physicochemical properties was of

necessary consequently. On the other hand, co-digestion of sewage sludge with

readily biodegradable substrates could enhance their anaerobic biodegradation

although to date there have been very little evidence to substantiate this hypothesis.

1.2 Scope of study

The increasing demand of renewable source of energy and adequate quality of

biosolids has provoked a great deal of work to formulate the feasible treatment

processes applied in WWTPs. In addition to sewage sludge stabilization, AD has

been known to produce biogas, which is a renewable fuel. Using organic materials as

co-substrate of sewage sludge is expected to enhance the efficiency of anaerobic

digesters. Furthermore, a more comprehensive understanding of key physiochemical

properties of the substrate, operational conditions, and biogas potential is of great

necessary prior to any large-scale operations.

The accumulation of TrOCs in biosolids is also a great challenge because they

severely affect the suitability of biosolids for land applications. Although the removal

of TrOCs by aerobic treatment has been extensively studied, little attention has been

paid to the anaerobic treatment. The assessment of their fate throughout the anaerobic

fermentation by batch-mode biomethane potential (BMP) tests can provide valuable

information on elimination capacities at different redox conditions.

1.3 Research objectives

BMP batch tests are conducted to evaluate the potential of the AD process in

producing biogas and treating TrOCs in sewage sludge. The outcomes will then

facilitate the performance of anaerobic digesters in water utilities in Australia and

around the world. The specific objectives of this research are:

Characterise sewage sludge and glycerol for parameters relating to the AD

process;

Assess the potential of methane (CH4) production of each sewage sludge and

glycerol;

Evaluate the generation of biogas during anaerobic co-digestion of sewage sludge

with glycerol; and

4

Evaluate the removal of TrOCs accumulated in sewage sludge from wastewater

treatment processes by the AD process.

1.4 Expected outcomes

The main expected outcome of this study is the better performance in terms of CH4

yield when co-digesting sewage sludge with other organic waste materials in lab-

scale batch experiments. Prior to this, data regarding the varied compositions of

sewage sludge and other organic waste materials will be identified; and the BMP of

each substrate is assessed. Another important outcome will be the extent of

elimination of some TrOCs by the anaerobic conversion.

1.5 Thesis outline

This thesis consists of seven chapters systematically shown in Figure 1. Chapter 1

gives an introduction of background and research objectives. Chapter 2 is a thorough

and updated review on the literature of AD and its applications on the co-digestion

together with publications of the fate of certain persistent compounds. Chapter 3

describes the research methods to fulfil specific objectives clarified in Chapter 1.

From Chapter 4 to Chapter 6, experimental results and data analysis are presented.

Chapter 7 will draw conclusions of this whole study and recommendations for future

studies.

5

Figure 1.The descriptive structure of thesis.

6

2 LITERATURE REVIEW

This chapter provides an overview of the current knowledge regarding anaerobic co-

digestion of sewage sludge and other organic waste materials. The AD process of

sewage sludge is firstly presented and discussed. This is followed by a

comprehensive review of CH4 production by anaerobically co-digesting sewage

sludge with other substrates. In addition, the behaviour of TrOCs during AD is also

summarised.

2.1 Sewage sludge

In the effort to improve effluent quality, WWTPs are built and upgraded. While these

plants can produce high effluent quality, sludge disposal remains an underlying issue.

These include the high cost of sludge treatment, which makes up more than 50% of

total wastewater treatment cost [28], and potential risks associated with sludge

disposal for the environment and human health [29].

Sewage sludge is a mixture of solids and semi-solids removed from the liquid stream

of WWTPs [30]. A more restricted definition is “a residual solid from sewage plants

treating domestic and urban wastewaters and from other sewage plants treating

wastewater of a composition similar to domestic and urban wastewaters” [10].

2.1.1 Types of sewage sludge

To assess options for sludge treatment and disposal, it is necessary to investigate

different kinds of sludges and their origins. A typical sewage treatment plant includes

preliminary, primary, secondary and tertiary processes [2]. During preliminary

treatment, large debris (sticks and papers for example), are removed and these do not

add to sewage sludge. On the other hand, residues from all the other processes are

collected as sludge (Figure 2).

Primary sludge is from the primary treatment process (commonly known as

sedimentation), containing high total solids (TS) content. The characteristics of

primary sludge varies considerably depending on the initial compositions of

wastewater, the efficiency of primary sedimentation and the usage of chemicals in

sedimentation, like coagulant aids [31]. Primary sludge can consist of oil, grease,

vegetable materials, faecal materials, papers, sanitary and medical waste, kitchen

wastes, and a variety of pathogens [32].

7

Figure 2. Different types of sewage sludge (modified from [10, 33]).

Primary settler Secondary settler Tertiary treatment

Co-settle

sludge

Aeration

Tertiary

sludge Activated

sludge

Surplus activated

sludge

Digested

sludge/Biosolids

Primary sludge

Secondary settler Biological filter Primary settler

Co-settle sludge Humus sludge/Biosolids

Anaerobic or

aerobic

digester

8

Treatment processes such as activated sludge process, trickling filter and rotating

biological contactors result in humus sludge or biological sludge [30]. Humus sludge

is the settled product from soluble waste in the primary effluent. This is a mixture of

microorganisms: sloughed bacteria and fungus under living or dead remains. Humus

sludge has an earthy smell and dark brown in colour.

Humus sludge from biological aerated filters and their variations, which have

different types of biological media, share certain characteristics with activated

sludge. In practice, humus sludge is returned to co-settle with primary sludge in the

primary settler.

Activated sludge is removed from the activated sludge process. Main components of

activated sludge are flocculated and synthesised solids and microorganisms [10]. Due

to the rate of recycling and other factors, activated sludge has low TS (1%) with the

colour ranging from grey, dark brown to black.

In the tertiary treatment step, the resulting sludge is called tertiary sludge. It has

fractions in common with the secondary sludge, which remains in the effluent of the

secondary treatment step and removed in the tertiary step. This sludge is normally

transferred to primary tanks and co-settle with primary sludge due to its small

amount.

Digested sludge, known as biosolids, is the product of biological digestion. This

process can be performed in a reactor with or without the presence of oxygen,

corresponding to the anaerobic or aerobic digestion processes. Biosolids contain

nutrient [34] thus should be considered as a resource. They may also contain

pathogens, which must be carefully managed for public health protection. Biosolids

classification is based on contaminant and stabilization grades. Once these grades are

evaluated, the beneficial use of biosolids can be divided into three categories:

Unrestricted, Restricted and Not Suitable for Use [35].

With respect to the pathogens, biosolids are classified as either Class A or Class B

according to the final part 503 regulations [34]. Class A biosolids contain a small

amount of pathogens. As such, sewage sludge must undergo treatment methods,

including composting, drying, and heat treatment, thermophilic AD process to

achieve Class A standards. Class B biosolids have a small level of pathogens and less

stringent requirements of treatment. Technologies, such as heating, composting,

9

digestion and pH adjustment, are in use to reduce the pathogen level to the point that

protects public health and environment.

There are other kinds of sludges resulting from other treatment processes. One of

them is physicochemical sludge coming from the physicochemical treatment of

wastewater. It is composed of flocs produced by chemical treatments like coagulants

and flocculants [33].

Combination of different sludge types is commonly utilised in sludge treatment. This

could be elucidated with diverse characteristics and compositions of mixed sludge,

facilitating downstream treatment processes. Regarding AD, the composition of

sewage sludge is a mixture of primary and secondary sludge [36-38].

2.1.2 Constituents of sewage sludge

It is necessary to know characteristics of sewage sludge for its effective treatment

and disposal. In general, sludge may include volatile, organic solids, nutrients,

pathogens, metals, organic pollutants and water [28, 39]. Table 1 summarises some

qualitative analyses of sludge constituents from the literature.

TS content affects the ability of sludge transference in the sewerage system. The

higher amount of TS, the more difficulty sludge flow gets. Therefore, it is needed to

remain sludge in liquid state, which will make sludge flow easily from a vessel and

through pipes. Sewage sludge should be characterised for TS content prior to any

sludge treatment processes.

The value of TS content after treatment can change depending on different treatment

methods. After thickening, TS content of sludge will increase up to 9%; and reach 25

– 35% after mechanical dewatering [33].

The solid content of sludge has 59 – 88% of volatile solids (VS) on dry weight basis.

VS content mainly contains organic compounds of animal or plant origin. It is

defined as the mass of solid materials that can be lost through evaporation or

oxidation at 550 °C.VS is an important parameter of the odour problem of sludge;

thereupon, the reduction of VS is one of the main objectives in sludge treatment. A

series of treatment methods, including AD, aerobic digestion, composting and

incineration, are used to minimise the VS content [1]. AD can biologically convert

around50% of VS to biogas.

10

Table 1. Typical constituents of different types of sludge [2, 10].

Constituent Unit

Type of sludge

Untreated

primary

sludge

Digested

primary

sludge

Activated

sludge

Co-

settled

sludge

Total solids % 2.0 – 8.0 6.0 – 12.0 0.83 – 1.16 nd

Volatile solids % of TS 60 – 80 30 – 60 59 – 88 nd

Grease and fat

Ether soluble

Ether extract

% of TS

6 – 30

7 – 35

5 – 20

nd

nd

5 – 12

14.7

Protein % of TS 20 – 30 15 – 20 31 – 41 32

Total nitrogen

(TN) % of TS 1.5 – 4 1.6 – 6.0 2.4 – 5.0 3.5

Total

phosphorus % of TS 0.8 – 2.8 1.5 – 4.0 2.8 – 11.0 2.8

Potassium % of TS 0 – 1 0.0 – 3.0 0.5 – 0.7 0.2

Cellulose % of TS 8.0 – 15.0 8.0 – 15.0 nd nd

Iron % of TS 2.0 – 4.0 3.0 – 8.0 nd nd

Silica % of TS 15.0 – 20.0 10.0 – 20.0 nd nd

Alkalinity mgCaCO3/L 500 – 1500

2500 –

3500 580 – 1100 nd

Organic acids mg acetic

acids/L 200 – 2000 100 – 600

1100 –

1700 nd

pH 5.0 – 8.0 6.5 – 7.5 6.5 – 8.0 nd

nd = no data

Nitrogen exists in sludge in either inorganic forms such as an ammonium (NH4+) or

nitrate or a complex organic form. The concentrations of both inorganic and organic

nitrogen are dependent on the types of sludge and handling processes. In comparison

to organic nitrogen, inorganic nitrogen is reduced more easily by dewatering and

drying procedures. Some insoluble sludge constituents like phosphorous or calcium

remain in sludge, and are further decreased during dewatering. In contrast, potassium

and sodium, soluble constituents, will follow treated wastewater [10].

Numerous specific organic chemical contaminants are present in sewage sludge. Of

these compounds, trace organic contaminants such as pharmaceuticals and endocrine

disrupting chemicals (EDCs) are of concern since their applications may results in

greater risks to human and environment [32]. In a recent literature review, Harrison

et al. [40] examined available data on the presence of organic chemicals in sludge in

US. Data were reported to 516 potential organic pollutants, which were grouped into

15 classes and their range of concentration was also recorded.

11

Raw sewage sludge, which does not experience any treatment, is regarded as a

source of pathogens, including bacteria, viruses, fungi, yeast, and parasitic

microorganisms. They present a public health hazard, so their quantity should be

figured out. In reality, the quantity and types of pathogens vary depending on the

health status of a community.

2.2 Anaerobic sludge digestion

2.2.1 Fundamentals of anaerobic digestion

AD is a process in which organic matter can be biodegraded in the absence of

oxygen by a consortium of microorganisms. An important product of AD is biogas,

which mainly comprises CH4, carbon dioxide (CO2) and traces of other gases [3].

AD involved a series of biochemical reactions, which can be divided into four stages,

namely hydrolysis, acidogenesis, acetogenesis and methanogenesis (Figure 3). AD

has been extensively used to treat biodegradable organics and produce biogas [5].

AD is a sequential process involving several complex biochemical stages. Each stage

is consistently performed under activities of a specific group of bacteria or

interactions of different ones. In the hydrolysis step, hydrolytic microorganisms

hydrolyse polymer materials to form monomers, such as amino acids and glucose.

These monomers are subsequently converted into H2, CO2 and short-chain fatty acids

such as acetic, propionic, and butyric acids in the acidogenesis step. In the

acetogenesisstep, syntrophicacetogenic bacteria metabolize these volatile fatty acids

(VFAs) to produce precursors for the methanogenic fermentation. Finally, CH4 is

formed from either acetate or CO2 and H2 by methanogenic bacteria in the

methanogenesis step [4, 41, 42].

2.2.2 Current status of the anaerobic digestion process

Anaerobic conversion is a natural process occurring in various environments, such as

wetlands, rice fields, intestinal tracts of animals, marine or fresh water sediments.

Humans have harnessed this process to take benefits as energy, rapid decomposition

of organic waste, and stabilised residue for a long time. Historically, applications of

AD could be dated back to the 10th

century, when it was used to produce combustible

gases from sediments in lakes, ponds, and streams. The very first design for AD with

the full-scale application was conducted for domestic wastewater treatment in the

12

18th

century. The recovery of biogas from well-designed sewage treatment facilities

was established in 1895 in England, treating 230 m3/d of wastewater; and the

collected gas was used to fuel street lamps [41, 43].

Figure 3. Different stages of AD process (modified from [4, 41, 42]).

Over the last two decades, a great deal of the literature has been published on

feasible applications of AD for solid waste and wastewater treatment. Apart from

biogas production, AD proves much greater potential due to more intrinsic merits,

including energy saving, nutrient recovery, reduction of waste organic content and

pathogen populations as opposed to the conventional aerobic digestion [4-6]. As a

result, extensive applications of AD have been only revealed very recently with a

number of developing designs by focusing on more complicated devices and

operational techniques, and increased understandings of microbiology and

biochemistry. There are various anaerobic reactor types in practice, of which batch

Complex particulate

organic matter

Soluble organics (simple

sugars, alcohol, organic acids,

ketones, amino acids)

VFAs

Hydrolysis

Acidogenesis

Acetogenesis

CH4

H2, CO2 Acetate

Hydrogenotrophic

methanogenesis

Aceticlastic

methanogenesis

13

reactors are the simplest configuration. The one-stage continuously fed systems, the

two-stage and multistage continuously fed systems were more advanced reactors

applied for AD treatment [44].

The evolution of AD applications was also confirmed by a broad range of potential

substrates for this process. Anaerobic technology such as single-phase (conventional)

and two-phase anaerobic digesters was often used in the treatment of dairy

wastewater for high energy production and waste stabilization. In terms of low

content of suspended solids in dairy wastewater, the conventional anaerobic reactors

are generally nominated for treatment. Currently, numerous studies of dairy

wastewater treatment have shown a wide range of applicable anaerobic reactor

designs, including downflow fixed film, anaerobic filter, up-flow anaerobic sludge

blanket (UASB), hybrid UASB, anaerobic sequencing batch reactors (ASBR), ASBR

upflow packed-bed, rotating biological contact reactor, upflow anaerobic solid

removal reactor, multichamber bioreactor, continuously stirred tank reactor (CSTR).

At laboratory scale, the efficient removal of chemical oxygen demand (COD) of

these reactors could reach up more than 90% [6].The authors also reviewed the two-

phase anaerobic treatment, which was employed for dairy wastewater consisting of

high concentrations of non-filtered solids and lipids. The dominant reactor type was

CSTR and upflow filter, with CSTR used for acidogenesis stage and upflow filter

responsible for methanogenesis. Compared to conventional AD processes, the two-

phase shows better outcomes with various kinds of industrial wastewater. Sludge

produced from WWTPs has also aroused much consideration since some strict rules

of sludge disposals were adopted [41]. Due to advantages of AD, it has become one

of the promising solutions for sludge stabilisation and energy production [30, 45, 46].

Current improvements of high-rate anaerobic system have been drawing more

attention on AD performances in agricultural waste treatment, especially animal

residues [44] which have different characteristics from those of municipal and

industrial wastewater [7]. Anaerobic treatment of the poultry and livestock manure

waste, two other types of agricultural waste, were also of interest due to increasing

concerns of their disposal [47-49], there has been having more and more

investigations of the AD process on them. The type of reactors used for livestock

manure waste treatment include: batch, continuous one stage, and continuous two

stage reactors, tubular reactor, ASBR, AF or the plug flow reactor (PFR) as

14

described in Table 2. Many studies have also indicated that co-digestion of poultry

and swine manures with each other or with additional substrates like sewage sludge,

some industrial byproducts proves more beneficial in biogas production and even

more efficient in substrate treatment [5, 44, 47, 50].

2.2.3 Factors influencing the anaerobic digestion process

AD can be sensitive to several operating factors, including pH, alkalinity,

temperature, and characteristics of the substrates (Figure 4). To optimise the

efficiency of AD, these factors should be carefully regulated.

Figure 4. Factors influencing AD performance.

2.2.3.1 pH

pH fluctuation can influence biogas yield throughout AD. In the early stages (i.e.

hydrolysis, acidogenesis, and acetogenesis), pH decreases due to the formation of

organic acids. Once the methanogenesis step occurs, pH may increase slightly

because of the production of ammonia [42]. Below pH 6, inhibition of CH4-forming

bacteria can occur and the anaerobic process can be disrupted [8]. The pH inside

digesters is an important factor governing the growth of anaerobic microbes,

particularly methanogens, through its impact on enzyme activities. This is because

each group of microorganisms has its own appropriate pH for growth. Methanogenic

bacteria are more sensitive to pH and need a pH range between 6.5 and 7.8 [47]

while acid-forming bacteria can function in a wider pH range from 4.0 to 8.5 [51] but

prefer a pH of 5.5 to 6.5 [44, 52]. In operation, it is necessary to keep the pH close to

neutral since methanogenesis is the yield-limiting step. Lime addition is a common

technique to overcome pH reduction.

AD process

pH

Alkalinity

Temperature

C/N ratio

VFAs

Substrate Operating factors

15

Table 2. Operational data of different anaerobic bioreactors used for different types of agricultural waste [5, 44, 47, 50].

Agricultural

waste Reactor configuration

Reactor

volume OLR

HRT

(days)

Temperature

(°C)

COD or

VS removal

Biogas or CH4

yield

Poultry manure

Batch reactor, UASB,

pilot and full-scale

digesters

0.1 L – 95

m3

1.1 – 2.9

gCOD/L.d 0.5 – 305 25 – 55

32 – 78%

COD

31 – 548

mLCH4/gVS

3.6 – 368

mLbiogas/gVS

Cattle manure

Fixed-film reactor,

attached-film bioreactor,

anaerobic rotating

biological reactor, batch

reactors, downflow

anaerobic filter, fixed

dome plant, UASB,

CSTR, UAF, TPAD,

AHR, and two-stage

anaerobic systems

120 mL -

1300 m3

0.117 – 7.3

gVS/L.d 0.5 – 140

5 – 82

37.9 - 94%

COD

7.3 - 92%

VS

93 – 382

mLCH4/gVS

103 – 450

mLbiogas/gVS

Swine manure

Hybrid UASB, CSTR,

baffled, ASBR, batch

reactor, dispersed growth,

stirred batch, and PFR

125 mL –

565 L

0.9 – 15.4

gVS/L.d 0.9 – 113 22.6 – 60

57 – 78%

COD

22 – 360

mLCH4/gVS

207 – 249

mLbiogas/gVS

Organic loading rate (OLR); hydraulic retention time (HRT); up-flow anaerobic filter (UAF); continuously stirred tank reactor (CSTR);

anaerobic sequencing batch reactor (ASBR); anaerobic hybrid reactor (AHR); temperature-phased AD (TPAD); plug flow reactor (PFR)

16

2.2.3.2 Alkalinity

Alkalinity refers to the buffering capacity, which is important for regulating pH in

AD. Alkalinity originates from the degradation of organics in the form of CO2,

bicarbonate and ammonia [51]. The equilibrium of CO2 and bicarbonate will resist

the rapid changes in pH. Compared to pH, alkalinity or buffering capacity gives more

reliability for system stability since the possible accumulation of VFAs can lead to a

reduction in buffering capacity and pH [53].

The pH in an anaerobic system is controlled by CO2 in the gas phase and bicarbonate

in the liquid phase. Thereby, pH will increase when more bicarbonate is added. In

practice, when the digester pH decreases a net strong base, either sodium hydroxide

or calcium hydroxide [47] or carbonate salts, are utilised. They are able to remove

CO2 in the gas phase to convert into bicarbonate. Bicarbonate can be directly added

to eliminate the lag time and over organic dosing [44].

2.2.3.3 Temperature

AD strongly depends on temperature because not only does temperature affect the

physicochemical properties of substrates in digesters, but bacteria are also sensitive

to any changes in temperature. Therefore, it is essential to maintain constant

favourable temperatures for the growth of anaerobic microbes [8]. Insulation, water

baths or passive solar heating are used for temperature maintenance; and heat can be

added by using heat exchanges in the recycled slurry or heating coils or steam

injection in the digester [13]. Any fluctuation of temperature even small change from

30 to 32 °C [44], would lead to inactivation of bacteria, resulting in a decrease in

biogas production. Moreover, process failure can be observed at temperature changes

in excess of 1 °C per day [54].

There are three temperature ranges investigated for applications: psychrophilic

temperature from 10 to 20 °C, mesophilic temperature from 20 to 40 °C, and

thermophilic temperature between 40 and 60 °C [47]. Once sufficient retention time

for the CH4-producing bacteria is provided, anaerobic sludge digestion could be

operated successfully at psychrophilic temperature as low as 20 °C [44]. The main

discrepancy between mesophilic and thermophilic digestion lies in the CH4 yield. It

is documented that higher CH4 produced by thermophilic digestion compared to that

by mesophilic digestion in a given digester due to the fact that high temperature is

17

the preferred condition for methanogens growth [8, 55]. Another advantage of

thermophilic digestion is pathogen removal since certain pathogens could be killed at

high temperature. Moreover, thermophilic conditions were evaluated to facilitate the

balanced fermentation system in biogas production by [56]. The application of high

temperature, however, has some drawbacks, such as increase in free ammonia or

VFAs, which make the process more susceptible to inhibitors [54].

2.2.3.4 VFAs

VFAs created during AD are an important intermediate product and relates to the

imbalance of AD. High VFA concentration primarily causes the process failures with

respect to an imbalance among acidogenic, acetogenic and methanogenic organisms

[14]. Limiting concentration of VFAs for a stable performance of a anaerobic

digester was reported at13000 mg/L by Viéitez et al. [57]. Additionally, less effective

removal of COD is observed with increased VFA production [47]. In the acetogenic

stage, the VFA accumulation will lead to a decrease in pH, which directly inhibits

the growth of methanogens. If inhibition lasts in long time, acetogens will

predominate in digesters. As discussed, the addition of buffering is an effective

solution since this can resist pH drop and maintain sufficient VFA concentration for

subsequent reactions [44]. While acetic acid is the key substrate for methanogenesis,

it is determined that propionic and butyric acids are inhibitory to methanogenic

bacteria. So as to avoid digester failures, monitoring of VFA, especially butyric

acids, has been shown to stabilise the overall system [44].

2.2.3.5 Ammonia

The present of ammonia in digester results from the breakdown of nitrogen-

containing matter, mainly from protein and urea. Ammonia is regarded as one of

inhibitory substances to the AD process [58]. Between two forms of ammonia, NH4+

and free ammonia (NH3) in liquid, free ammonia has been identified as the main

cause of inhibition. The reason is the hydrophobic form of ammonia could easily

penetrate through cell walls, causing pH imbalance and enzyme malfunction [58].

This inhibition in general has been clearly observed in the methanogenesis stage.

Koster and Lettinga showed that along with the increase of ammonia concentrations

in the range of 4051 – 5734 mgNH3–N/L, acidogenic populations in the granular

sludge were hardly affected while the methanogenic population lost 56.5% of its

18

activity [59]. On basis of CH4 production, ammonia has stronger effect on

aceticlastic than hydrogenotrophic methanogens. It is suggested that the free

ammonia should be kept below 80 mg/L, while ammonium could reach up to 1500

mg/L without causing any inhibition [55]. It is pH and temperature that become

factors influencing the ammonia inhibition capacity through ammonia concentrations

[58]. The more pH increases, the more the amount of ammonia and less the amount

of ammonium are.

2.2.3.6 C/N ratio

The C/N ratio is a common parameter that has been thoroughly investigated by

numerous studies of AD. The C/N ratio represents the relative amount of organic

carbon measured by COD and nitrogen present in feedstock. The changes in the

specific CH4 yields and CH4 content consistently comply with the C/N ratio. Low

nitrogen content would lead to the inhibition of AD since anaerobic microbes needs

an adequate amount of nitrogen for their growth while organic carbon is considered

as a sole source for anaerobic activity. Increase in pH can result from low C/N ratio.

On the other hand, a high C/N ratio would directly result in a rapid conversion of

nitrogen and low biogas production. It has been elucidated that the ideal levels of

C/N ratio should be in the range of 20 – 30, typically approximately 25 [13]. For

instance, the peak values of CH4 yields were observed at a C/N ratio of 23 [60].

2.2.4 Anaerobic digestion in sewage sludge treatment

Through wastewater treatment systems, only liquid can be disposed of to the

environment while solids are collected for further treatment before discharge, of

which, sludge is by far the largest component in volume. Sludge is considered as

both a potential pollutant and a prospective source. Accordingly, sludge treatment

and disposal should be always an integrated part of wastewater treatment. Therefore,

it is imperative to seek environmentally sound and sustainable methods for sewage

sludge handling and disposal. Sustainable sludge treatment may be defined as “a

method that meets requirements of efficient recycling of resources without supply of

harmful substances to human or environment” [2]. Thereby, not only sludge

management but also research into innovative handling methods should concentrate

on three main objectives: (1) complete sludge treatment with regards to reduction of

sludge levels, (2) recovery of valuable products from sludge, regarding the nutrient

19

used as a soil conditioner or improver, and biogas production, and (3) choice of a

treatment method at an acceptable cost [28, 31].

In terms of disposal of excess sludge, the 1990 North Sea Conference consented to

an agreement of banning of the sludge dumping at sea and instead using biosolids in

agriculture. There is a need to find replacements for the dumping of sludge at sea.

The possible fate of excess sludge are landfill, incineration, spray irrigation, drying,

composting, and AD. They are towards the final goal – the transformation of

wastewater sludge into the innocuous and easily dewatered form. Several studies [54,

55, 61] have confirmed that AD is an ideal method compared to other methods of

sludge destabilisation based on its merits as shown in Table 3. This table also lists

some drawbacks of the anaerobic sludge digestion, which greatly demands

consideration.

Table 3. Advantages and disadvantages of anaerobic sludge digestion [5, 6, 10, 52,

62].

Advantages

- Production of CH4, a renewable source of energy, compensating energy for

maintaining the temperature of digesting sludge, and meeting requirements for

mixing; additionally, heat buildings, drive engines of aeration blowers or

generate electricity that can be used to run the sewage pumps

- Net reduction of mass and volume of sewage sludge during the conversion of

organics into CH4 and CO2 gas and water

- Transformation of solid content to stable, inoculum sludge

- Beneficial reuse in agriculture – soil conditioner or fertilizers: treated sludge

may contain N, P and other nutrients and organics can improve soil quality

- Inactivation of pathogens during AD

Disadvantages

- The high capital costs: large, covered tank as well as pumps for feeding and

circulating sludge, heat exchangers and compressor for gas mixing

- Long HRT necessary to develop and maintain the population of CH4-

producing bacteria

- The quality characteristics of supernatant from anaerobic sludge digestion are

poor, containing suspended solids, dissolved and particulate materials,

nitrogen and phosphorus

20

2.3 Anaerobic co-digestion of sewage sludge with other substrates

The efficiency of AD can be determined as the volume of produced biogas or the

amount of substrate depletion or the formation of intermediates and other final

products during. These performance indicators are inter-related, thus, it is generally

reliable to assess the performance of AD based on the biogas yield. The more

efficient the anaerobic treatment, the more biogas/CH4 production will be generated.

Biogas produced from AD facilitates a sustainable development of energy supply

both economically and environmentally. As compared to the energy yield of aerobic

degradation, anaerobic biomass conversion results in low energy, because which is

mainly stored in biogas. This energy is subsequently harvested in the presence of

oxygen by aerobic organisms or by humans through heating and other processes [63].

2.3.1 Advantages of anaerobic co-digestion in treating sewage sludge

With the aim of improving biogas quality and quantity, certain possible approaches,

including process design improvement, pretreatment of substrates, removal of toxic

components and co-digestion should be considered. Improving process design could

be performed via either increasing biomass retention since a dense mixture of

microorganisms is essential for the biochemical metabolism of complex substrates,

or advancing configuration and operation in terms of temperature and HRT. As an

obvious example, enhanced biogas production levels and COD/VS removal

efficiency of manure along with a series of bioreactor designs were reported (Table

2). Regarding sludge, Athanasoulia et al. [57] compared biogas production potentials

between a cascade of two methanogenic CSTR in a series, and a conventional one-

step CSTR reactor treating sewage sludge, concluding that the biogas production was

considerably improved by 9.5% to 40.1% in the serial digestion. In terms of VFAs,

values were higher in the cascade configuration, from 31.5% to 33.8%, in

comparison with the one-step process, from 36.2% to 40.7%. Another way to

increase biodegradability is the pretreatment of substrates which are hardly

hydrolysed with a high content of cellulose and hemicellulose [64]. Mechanical

destruction, heat treatment, and chemical treatment are pretreatment that techniques

have been extensively applied for domestic sludge [14, 46, 65, 66]. The removal of

toxic or inhibitory compounds in the feed prior to AD, contribute to increase biogas

yield. Ammonia levels can be decreased by free-ammonia stripping or bentonite

21

bound oil while ferric chloride and ferrous chloride pose their effect on precipitating

sulphate/sulphide [14, 67]. Up to now, results from the literature have clearly

demonstrated that the most successful way in improving biogas yield should be co-

digestion [68-71]. This can be explained by certain benefits of co-digestion: (1)

dilution of potential toxic compounds, (2) the supply of missing nutrients, (3)

synergistic effects of microorganisms, (4) increased load of biodegradable organic

matter, (5) economic advantage of sharing equipment, and (6) better biogas yield [16,

17]. AD has mainly taken place for the treatment of sewage sludge or other wastes.

Recently, for instance, manure from pigs, cows, and chicken is currently digested

with co-substrates, which include harvest residues, organic wastes, food wastes,

municipal bio-wastes, and energy crop for higher biogas production [8, 9].

2.3.2 Current status of anaerobic co-digestion of sewage sludge and other organic

materials

Although co-digestion could offer many benefits [69, 72], the application of co-

digestion of sewage sludge and other organic materials is still not well understood. In

recent years, much successful efforts have been made for such co-digestion in order

to upgrade the role of anaerobic degradation in stabilising sewage sludge and

produce feasible bioenergy as well [73].

The quality and composition of biogas produced could not be significantly changed

by OLR, HRT or other operational parameters but considerably depend on their

origin and substrate compositions, such as carbohydrates, protein, and lipids [74],

whose concentrations in sewage sludge were documented to be varied from one type

of sludge to another one (Table 1). Among them, lipids has been known as a very

promising substrate with regards of the CH4 production (i.e. 1014 mLCH4/gVS), but

requires more time for complete biodegradation. Meanwhile, proteins and

carbohydrates show faster conversion rate but lower biogas levels, i.e. 496

mLCH4/gVS and 415 mLCH4/gVS, respectively [8, 64, 75]. Through investigations

on 175 BMP assays, Labatut et al. [76] revealed that substrates rich in lipids and

easily biodegradable carbohydrates yield the highestCH4 potential, while others are

more recalcitrant with a high lignocellulosic fraction and pose the lowest gas

production rates. This could explain why hydrolysis has been reported as the rate-

limiting step in the AD of sewage sludge since proteins have been reported as the

22

rich composition in sewage sludge rather than carbohydrates and lipids. Therefore,

the addition of easily hydrolysed substrates to sewage sludge during AD must

initially accelerate the growth of microorganisms, speeding up the whole process as a

result. One of the key operating parameters of AD process, the C/N ratio, is another

concern related to sewage sludge. As discussed above, due to the low C/N, around 6

– 9, co-digestion of sewage sludge with high carbon content substrates is a necessity

for better anaerobic degradation and higher biogas production.

Taking the biogas production potential into account, any type of biomass containing

carbohydrates, proteins, fats, cellulose, and hemicelluloses as main components

could be typical substrates for AD [77]. As such, potential co-substrates for AD

should be categorised into harvest residues, including top and leaves of sugar beets,

organic wastes, food wastes, municipal biowaste from households, energy crop, and

sewage sludge [8, 9]. Table 2 summarises the biogas/CH4 production potential and

VS removal ability when mixing sewage sludge with some different promising

substrates, including fats, oil and grease (FOG), animal byproducts from the meat-

processing industry, and coffee waste, organic fraction from municipal solid wastes

(OFMSW), brewery sludge, manure, and glycerol. Anaerobic co-digestion has been

mainly performed at mesophilic temperature range, from 35 °C to 37 °C, with

various bioreactor systems. Operating at thermophilic temperature lowered the CH4

yield but enhanced VS removal. It is clear that the addition of co-substrates makes

AD achieve greater biogas production. The current studies have shown the

differences in biogas percentages in a mixture of biogas. In comparison, the highest

CH4 content was observed in the gas coming from the sewage digester, from 57.8%

to 65%, while the lowest CH4 and highest nitrogen content were found in the landfill

gas, from 37% to 62% [78-81]. From Table 4 of typical concentrations of biogases, it

can be seen that CH4 and CO2 are the predominant components resulting from

anaerobic degradation. Unlike CO2 which is more water soluble, CH4 is partially

non-soluble and takes up for the most part in gas phase [68]. Therefore, it might be

more reliable to assess CH4 potential rather than biogas potential.

Each co-substrate has its own CH4 production capacity. Sewage sludge, typically

primary sludge and waste activated sludge [70, 72], contains certain easily

degradable material [16]; and its CH4 production potential could be approximately

300 – 400 mLCH4/gVSadded [69, 71].

23

Table 4. Biogas composition [9].

Combustible ingredients Concentration (%)

CH4

H2

H2S

50 – 70

< 1

2

Non-combustible ingredients Concentrations (%)

CO2

H2O

O2

NH3

25 – 50

2 – 7

0 – 0.5

0 – 2

As mentioned above, lipids are the most attractive macromolecular component

during the anaerobic transformation for its high theoretical CH4 production.

Luostarinen et al. [82] reported that the CH4 yield potential of grease trap sludge

(GTW) is 918 mLCH4/gVSadded. The chemical and physical properties of GTW could

differ greatly depending on origin and grease abatement device (Table 5). Several

studies on this process observed an increase of 30% in biogas production and a

recovery of more than 50% in energy for on-site generation when directly adding

FOG, the top floatable layer of GTW, from the food industry [83, 84]. Although

FOG fraction in GTW is low, at average value of around 2 – 3% by volume [85], it

has been considered a potential feedstock for both AD and biodiesel production

processes due to its high fraction of lipids [72, 86]. However, lower pretreatment

expenses during the anaerobic FOG conversion makes it a more feasible option of

disposal compared to the biodiesel production [85].

In recent years, despite enhanced CH4 potential has been achieved by anaerobic co-

digestion of FOG and sewage sludge, several operational concerns of the inhibition

of methanogens due to FOG and its derivations have been stated. The detrimental

effects of FOG on methanogenic bacteria greatly attributed to the high content of

long chain fatty acids (LCFAs). There are three main mechanisms, including the

inhibition of aceticlastic and hydrogenotrophic methanogens due to the LCFAs

toxicity, the sludge flotation and biomass washout due to the adsorption of LCFAs

onto sludge, and the movement limitation of bacteria when its cell wall is covered by

a LCFAs layer [85-87]. Therefore, the AD process of FOG starts slowly. The CH4

production of FOG was only improved when FOG is mixed with sewage sludge. The

higher CH4 yield achieved from this digestion is illustrated in Table 6.

24

OFMSW is also a suitable co-substrate since it is the easily degradable content of up

to 40% in municipal solid wastes organic fraction, accounting for more than 40% of

municipal solid wastes (MSW) [88], is the easily degradable constituent. Concerns

related to the OFMSW management via conventional methods, include composting,

incineration, landfilling, and dumping [56]. Accordingly, anaerobic co-digestion of

OFMSW and sewage sludge becomes one valuable alternative to these approaches in

terms of energy recovery and environment security. Differences in characteristics of

OFMSW are caused by various sources. In one study carried out by Cabbai et al.

[89], source selected OFMSW from canteens, bakery, restaurants, supermarkets, and

households vary in compositions (Table 5). Compared to sewage sludge, OFMSW

has a higher C/N ratio and lipid load that elevate the initial stage of AD, which may

benefit CH4 production in case of co-digestion [88]. Significantly, higher CH4

productivity during co-digestion was observed under various conditions compared to

the single sludge anaerobic conversion (Table 6). Among several OFMSWs, food

waste in general is also considered the most variable substrate depending on its

origin, such as household, restaurants, markets, university dining hall, and

components, in terms of carbohydrates, proteins, and lipids [78]. Food waste

diversity in characteristics is also expressed in element compositions, in which

sufficient carbon content makes food waste the highly degradable substrates in AD.

In general, food waste has high VS/TS ratio (80 – 90%) and moisture content (75 –

85%), which makes landfill applications of food waste inadequate due to

environmental issues, namely odour release, toxic gas emission, and groundwater

contamination [90].

Instant coffee waste characteristics are mainly dependent on the production

procedure applied and raw matter used. In general, coffee waste mainly contains

carbohydrate in forms of fibres, namely lignin, cellulose, and hemi-cellulose

regardless of original materials used [91, 92]. On one hand, it has been reported from

the literature that AD of coffee waste was implemented at both mesophilic [92] and

thermophilic temperatures [35] for such high carbon content. These investigations,

on the other hand, indicated that the high concentration of solids and fibre could lead

to some problems such as clogging, but still yield the high CH4 composition of 70%.

This makes necessary to evaluate the AD performance of each kind of waste in an

anaerobic co-digestion with other substrates. The feasibility of anaerobic treatment of

25

coffee and sewage sludge was assessed, and resulting CH4 yield together with VS

reduction were shown relatively high (Table 6). The low CH4 yield of manure is

attributed to the high water content as well as high content of fibre, typical ranging

from 10 to 20 m3

of CH4 per tonne of treated manure [93].

Table 5. Characteristics of certain co-substrates used for anaerobic digesting with

sewage sludge.

Types of co-substrates Crude glycerol FOG OFMSW

Char

acte

rist

ics

TS (%) nd 1.8 – 97.2 8 – 70

VS/TS (%) nd 88.9 – 98.6 84.1 – 98.0

COD (g/L) 1, 054 – 1,216 nd nd

pH 5.0 ± 0.1 nd 3.7 – 4.6

Density (kg/L) 1.25 ± 0.1 nd nd

Electrical conductivity

(μS/cm)

4.2 ± 0.3 nd nd

TN (% of TS) nd nd 1.7 – 3.3

C/N nd nd 13.7 – 31.4

Com

ponen

ts

(% w

/w)

Carbohydrates nd 0.5 – 15 10.7 – 42.5

Protein nd 0.3 – 7 9.7 – 20.5

Fat nd 78 – 99.5 5.8 – 25.0

Glycerol 50 – 60 nd nd

Alkalis 12 – 16 nd nd

Methyl esters 15 – 18 nd nd

Methanol 8 – 12 nd nd

References [18-20] [85] [89]

nd = no data

Considered as a non-toxic, affordable, renewable and environmentally friendly

alternative to petroleum-derived fuels, biofuels, namely biodiesel and bioethanol,

have become the greatly demanded energy source. This led to the exponentially

increasing production of biofuels along with a surplus of crude glycerol since

glycerol is an inevitable co-product from not only biodiesel chemically and

enzymatically manufacturing [94], but bioethanol production also [95, 96]. Crude

glycerol is the main co-product in the biodiesel production. The ratio of produced

crude glycerol and biodiesel was estimated at 0.1 [18]. Consequently, such extra

crude glycerol has generated both a reduction in its market price, ten times as

reported recently by Yazdani et al. [95] and an environmental issue due to the

disposal of waste stream containing glycerol [18] as well as polluted wastewater

26

from purification processes [20]. There was a need of effective treatment processes

to transfer crude glycerol to higher valuable products that could improve economic

feasibility of biofuel industry. CH4 production by the anaerobic fermentation of

glycerol offers a great chance for energy supplementation in biofuel industry. It is the

high carbon content of crude glycerol that makes it an ideal substrate for anaerobic

conversion when the high production of biogas is taken into consideration.

Advantages of crude glycerol are the highly solubility since major components of are

pure glycerol, alkalis, methyl esters, and methanol as given in Table 5, as well as the

easy storage capacity of glycerol over a long time. It was reported that a large range

but low concentration of impurities, especially elements, caused by glycerol

purification processes and different used feedstocks have not exerted any significant

inhibitions to AD. The high salinity, sodium and potassium salts of crude glycerol

could have a negative effect on microbial activity [18]. The failure of the anaerobic

degradation of crude glycerol possibly results from the low concentration of N,

which is an essential element for microorganisms. However, the co-digestion with

sewage sludge completely turns it beneficial due to low C/N of sewage sludge. Until

now, little research work has been published regarding such co-metabolism. When

treating sewage sludge from WWTPs, Fountoulakis et al. [19] investigated that the

supplementation of glycerol can increase biogas production (Table 6) at

concentration of 1% by volume. Imbalance of the system was observed when adding

higher glycerol concentrations.

27

Table 6. Co-digestion of sewage sludge with different substrates.

Feedstock Bioreactor system Temperature

(°C) CH4/Biogas production

CH4/biogas

increase (%)

VS removal

(%) References

SS + FOG

(52 : 48) Semi-continuous

digester

35 449 mLCH4STP/gVSadded 195 45

[84] 52 512 mLCH4STP/gVSadded 160 51.2

SS 35 159 mLCH4STP/gVSadded - 25.2

52 197 mLCH4STP/gVSadded - 30.7

SS + GS (90 :

10) Continuous-pilot scale

digester 35

295 NmLCH4/gVSadded 9

45 – 58 [37] SS + GS (70 :

30) 344 NmLCH4/gVSadded 27

SS 271 NmLCH4/gVSadded - nd

SS + GS (54 :

46) Lab-scale digester 35 463 ± 48 mLCH4/gVSadded 66.5 59 – 70 [82]

SS 278 mLCH4/gVSadded - nd

SS + FOG

(80 : 20) Lab-scale single stage

digester 35

510 mLbiogas/gVSadded 18 57

[38] SS + FOG

(40 : 60) 640 mLbiogas/gVSadded 50 59

SS 430 mLbiogas/gVSadded - nd

SS + OFMSW

(50 : 50) Fed-batch digester

35 360 mLCH4/gVSadded 93.5 57

[56] 55 180 mLCH4/gVSadded 45 64

SS 35 186 mLCH4/gVSadded - 26.6

55 124 mLCH4/gVSadded - nd

28

Feedstock Bioreactor system Temperature

(°C) CH4/Biogas production

CH4/biogas

increase (%)

VS removal

(%) References

AS + Coffee

waste

Dual digestion system 35, 55 310 mLCH4/gVS nd 41 – 52 [97] Single stage digester 35 280 mLCH4/gVS nd 28 – 50

AS + Coffee

waste Batch assay 37

240 – 280

mLCH4STP/gVSadded nd 75 - 80 [91]

SS + ABP Lab-scale digester 35

400 ± 30 mLCH4/gVS nd nd [98] SS + ABP 410 ± 30 mLCH4/gVS nd nd

SS + brewery

(75 : 25)

Bench-scale CSTR 36 ± 0.2

510 mLbiogas/gVSremoval 27.5 16.4

[99] SS + brewery

(50 : 50) 610 mLbiogas/gVSremoval 52.5 18.2

SS + brewery

(25 : 75) 650 mLbiogas/gVSremoval 62.5 19.0

SS 400 mLbiogas/gVSremoval - 14.9

SS + FW +CM

(10 : 20 : 70)

Continuously stirred-

tank reactors 36 603 mLCH4/gVSadded nd nd [100]

SS + Gly

(99 : 1) CSTR 35 2353 ± 94 mLCH4/d 112.7 nd [19]

SS 1106 ± 36 mLCH4/d - nd

Standard temperature (0 °C) and pressure (1 atm) (STP); normal mL at 1 °C and 1 atm (NmL); sewage sludge (SS); grease trap sludge (GS);

activated sludge (AS); animal byproducts (ABP); food waste (FW); glycerol (Gly); and cattle manure (CM); nd = no data

29

2.4 The behaviour of trace organic contaminants during the anaerobic

digestion process

Concerns about the occurrence of TrOCs in wastewater and sewage sludge have

increased in recent years. It have been widely reported that their concentration levels,

removal rates as well as their behaviour significantly differ throughout wastewater

treatments. This could be explained by a variety of applied treatment processes,

operational conditions, and other factors such as temperature, pH and seasonality

[21, 22]. During wastewater treatment, TrOCs will appear in the supernatant to be

recycled in outlet and partly accumulate in sewage sludge [22, 25]. Carballa et al.

[25] conducted a comprehensive study on the behaviour of pharmaceutical and

personal care products (PPCPs) during wastewater treatment and detected most of

PPCPs, concluding that they were present in both the primary and secondary sludge

fed to anaerobic digesters. Samaras et al. [21] determined that the removal of such

TrOCs as nonylphenol and triclosan was attributed to the absorption mechanism onto

sludge while biodegradation/biotransformation was the significant mechanism for

contaminants such as ibuprofen, diclofenac, and ketoprofen.

2.4.1 Trace organic compounds and their effects

Although the classification of TrOCs has not been fully published, these compounds

could be simply divided into certain main categories, including pesticides,

pharmaceutically active compounds, and EDCs. This classification mainly depends

on their particular intended usage. That means their source might correlate to where

they have been employed. Through several routes, TrOCs could contaminate natural

water bodies, such as lakes, rivers and underground streams. Numerous documents

have reported the potential effects of TrOCs on human health, aquatic biota, and

environment [101]. Plants and animals could take up some anthropogenic organic

chemicals inside sludge, which might accumulate in the food chain. As such, some of

them potentially have severe effects on flora and fauna and soil as well. These effects

pertain to blocking or hampering functions of hormones in human and animals [101],

feminisation of male fishes by EDCs [102], estrogenic effects in rates caused by

bisphenol A [103].

30

2.4.2 Occurrence of emerging contaminants in sludge

Concentrations of TrOCs range between few ng/L to some g/L in wastewater, surface

water, groundwater and drinking water. While in sewage sludge from WWTPs, their

concentrations could reach up to a few mg/kg [21].

Differences in concentration are observed among different groups of compounds as

well as different countries and WWTPs. This is because their concentrations in

sludge widely depend on various factors. Concentrations of original compounds in

influent wastewater directly govern their final concentrations in sludge. Another

factor is their physicochemical properties, such as solubility and molecular weight.

Sludge characteristics and operational conditions applied in WWTPs also have great

effects on occurrence of TrOCs in sludge [104, 105].

The abundance of TrOCs in sludge has been reported at different levels. In general,

sludge is dominated with surfactants, of which linear alkyl benzene sulphonates

(LAS) has the highest concentrations in sludge due to their wide usage. Other two

surfactants taken into account are nonylphenolethoxylates and quaternary

ammonium-based compounds. Concentrations from few mg/kg to more than g/kg has

been reported as a range of surfactants in sludge [106]. Regarding PPCPs, the highest

concentrations in sludge belong to hydrophobic compounds, including triclosan,

triclocarban, and galaxolide. Their range of concentration vary from some μg/kg to

mg/kg [107]. Some TrOCs from industrial applications, brominated flame retardants,

organotins, and perfluorinated compounds have been detected at lower

concentrations, few μg/kg [108-110] in sludge compared to phthalic acid esters, in

the range of some mg/kg [111] while limited data have reported in the presence of

nanoparticles and benzothiazoles.

In the sector of agriculture, several guidelines stated the limiting concentrations of

these compounds in biosolids management. Accordingly, their concentrations in

sludge vary depending on a particular contaminant. EU, for instance, proposed their

acceptable levels in sludge prior to agricultural reuse as shown in Table 7.

31

Table 7. Proposed limit values of TrOCs in sewage sludge [112-114].

No TrOCs Limiting concentrations in sludge

prior to reuse (μg/g)

1 Linear alkylbenzenesulfonates 2600

2 Diethylhexylphthalates 100

3 Nonylphenolethoxylates 50

4 Dioxins and furans 0.0001

5 Organohalogens 500

6 Polychlorobiphenyls 0.8

7 Polycyclic aromatic hydrocarbons 6

2.4.3 Anaerobic digestion in the removal of trace organic compounds

The appearance of TrOCs in sewage sludge has been widely reported in the

literature. Their concentrations considerably differ among different groups of

compounds, ranging from some μg/kg to several g/kg [22]; even for the same group,

various concentration levels are often observed between different countries and

WWTPs. The detection of target compounds in sewage sludge raised the public

consideration of their possible threats to the environment and human health since

biosolids containing TrOCs has been currently applied for agricultural purposes.

Considerable efforts have been made to assess the fate of TrOCs during the sludge

treatment process in general and anaerobic sludge digestion. A question remains

whether their occurrence in sewage sludge would become one of inhibitions or a

potential substrate for AD.

Until now, most reported data in the literature have indicated that some surfactants

and pharmaceuticals, generally antibiotics, pose an inhibitive effect on AD, while

only minor impact caused by other contaminants. In a study investigating the effect

of LAS on the anaerobic sludge digestion, a partial and total inhibition of

methanogenic activities was observed at high concentrations of LAS homologues

[115]. By using a mixed, mesophilic methanogenic culture, Tezel et al. [116] showed

that quaternary ammonium-based compounds were more inhibitory to

methanogenesis, at and above 25mg/L, than acidogens. Several anaerobic microbes

could be sensitive to certain antibiotics at a wide range of concentrations, influencing

the overall AD performance. For example, carbamazepine, sulfamethoxazole and

diclofenac were studied at mesophilic conditions with a wide range of concentrations

by Fountoulakis et al. [117]. The authors observed 50% inhibition on methanogenic

activity from 30 mg/L to more than 400 mg/L. A higher range of concentrations of

32

antibiotics, between 24 mg/L and more than 1,000 mg/L, showed moderate inhibition

effects [118].

Little and even conflicting studies on the behaviour of trace organics by AD have

been carried out although PPCPs in sewage sludge have been described to be one of

the most ubiquitous contaminants [25]. Some PPCPs have been reported to exhibit

some resistance to the anaerobic conversion process. In the liquid phase of digested

sludge, most available PPCPs were observed to be persistent [119]. The authors

stated that diclofenac offered severe inhibition to AD at high concentrations while

the two other PPCPs had no effect even at the high concentrations. Nonetheless,

removal efficiencies of PPCPs during this digestion have also documented by a

variety of comprehensive studies, varying from very high removal to nearly no

removal. Excellent PPCPs removal was confirmed for naproxen, sulfamethoxazole,

roxithromycin and oestrogens at more than 85%, followed by galaxolide, tonalide

and diclofenac at 60%. The value ranged between 20 and 60% with the exception of

carbamazepine, which showed no degradation [24, 25]. These eliminations achieved

were attributed to sludge adaptation to AD. In another study on the fate of

pharmaceuticals, ibuprofen and naproxen removal were once confirmed with

considerably percentages more than 80% [21].

Regarding surfactants, it is believed that AD of sludge could degrade some persistent

compounds. LAS is one particular example with different removal reported in

several studies [25, 120, 121]. More data relating to the mechanism and removal

efficiencies of nonylphenolethoxylates during the anaerobic conversion have been

reported. In general, nonylphenolethoxylates is biotransformed into shortened chain

formations such as nonylphenolpolyethoxylates and nonylphenol [122].

Nonylphenolpolyethoxylate in some studies were metabolized into nonylphenol [27].

Nonylphenol could be degraded during the anaerobic sludge digestion process to

some extent [23, 26, 27]. There is little literature assessing the biodegradation

kinetics of these compounds with the exception of nonylphenol, whose

biodegradation followed the first order kinetic model [26].

The most striking example of contradictory results for the fate during the anaerobic

sludge treatment is the group of estrogenic compounds. The rationale likely lies in

inherent difficulties of analysing these compounds in sludge matrix and their unclear

behaviours. From equivalent loads of estrogens in both inlet and outlet of digesters,

33

Andersen et al. [123] concluded that estrogens were not degraded under

methanogenic conditions. 17-β-estradiol-3-sulphate and estrogen-3-sulphate

concentrations were even increased by activities of strictly anaerobic desulphating

strains [124]. For estrone, 17-β-estradiol, and estriol, a significant extent of removal

efficiencies was presented under thermophilic condition and various sludge retention

times [24] or with different types of sludge under mesophilic and thermophilic

conditions [23] at laboratory scale. Nevertheless, some authors confirmed the

opposite. From equivalent loads of estrogens in both inlet and outlet of digesters,

Andersen et al. [123] concluded that estrogens were not degraded under

methanogenic conditions. Using mass balance, Muller et al. [125] indicated that their

elimination were not effective (ca 40%) at the plant scale. There is little data

reporting the biodegradation of 17-α-ethinylestradiol during the sludge AD process

[126]. Even the mass flow in the outlet for 17-α-ethinylestradiol was observed to be

higher compared to the inlet in one recent study [125].

In sewage sludge, phthalic acid esters were detected in sludge at higher

concentrations in comparison to certain compounds mainly originating from

industrial applications, such as organotins and brominated flame retardants [22]. The

mechanism of anaerobic biodegradation pertaining to phthalates has aroused greater

consideration to date. Some literature has presented different removal efficiencies of

phthalates during sludge AD. Dimethyl, diethyl and di-n-butyl phthalate esters were

degraded within one week [127], nearly completely in a lab-scale experiment [128,

129]. It should be mentioned that the most hardly removed phthalic acid ester is

bis(2-ethylhexyl) phthalate, even persistent during the anaerobic biodegradation

[127]. By using mass balance in a full-scale WWTP, Marttinen et al. [130] showed

that the percentage of bis(2-ethylhexyl) phthalate removal just reached around 32%

via AD of sewage sludge. Only when some operational parameters, such as

temperature and type of inoculum and HRT are improved, the elimination in some

extent of this compound was identified [131, 132].

Most studies on TrOCs removal during the anaerobic sludge digestion process were

operated at the full-scale WWTP. This is probably the greater demand of assessing

the fate of TrOCs in the discharge flow. Although their eliminations at laboratory

scale have been poorly studied, the greater efficiencies have been recognised for

many compounds, which were reported to be refractory [23, 24, 128]. Some practical

34

methods, such as pretreatment processes or a combination of AD with other

treatments could be applied during recent years. The co-digestion of sludge with

some kinds of readily biodegradable carbon source was also indicated to enhance the

anaerobic biodegradation [26, 121], limited research regarding this enhancement has

been reported to date though. This may be explained by improving the overall

mechanism through accelerating the hydrolysis step. For example, Chang et al. [26]

demonstrated this effectiveness on nonylphenol removal, whereas the optimised

elimination of polycyclic aromatic hydrocarbons was suggested to reach via the co-

metabolism by Barret et al. [112].

2.5 Summary

The AD process, a sequential-stage process with several complex biochemical

processes involved in the absence of oxygen, has aroused much interest due to its

advantages including the production of renewable energy in the form of CH4 and

lower installation cost compared to its counterpart, the aerobic digestion process. The

efficiency and instability of anaerobic sludge degradation greatly depend on several

operating conditions (pH and temperature for example) and properties of sewage

sludge. Higher performance in terms of CH4 production has been achieved

anaerobically co-digesting sewage sludge with various organic waste materials. Of

them, glycerol was a considerable co-substrate due to high in solubility and rich in

easily degradable organic components, little research has been in the literature to date

though. In respect of removal efficiency by the AD process, more consideration must

be given to TrOCs occurring in sludge because of their detrimental effects on

ecosystem. Little and even conflicting studies on the elimination of TrOCs under

anaerobic condition have been assessed at laboratory scale although these trace

contaminants determine the suitability of biosolids for agricultural applications.

35

3 MATERIALS AND METHODS

This chapter details experimental design, experimental protocols, and analytical

methods applied throughout this study. Materials, namely inoculum, substrate and

selected TrOCs and their properties were also introduced.

3.1 Materials

3.1.1 Inoculum

Digested sludge from a Sydney Water sewage treatment plant was used as inoculum.

This inoculum was withdrawn from a full-scale anaerobic digester operating under

mesophilic condition. The digester was continuously fed with only primary sludge,

thus, the collected inoculum has already acclimatised with raw primary sludge. The

inoculum was stored at 4 °C in the dark until utilisation.

3.1.2 Substrates

Raw primary sludge was also collected from a Sydney Water sewage treatment plant.

Pure glycerol (>99%) was obtained from VWR International (Australia). Two kinds

of crude glycerol were also obtained from the biodiesel industry (through Sydney

Water). They are identified as Biodiesel and BIA, respectively. All substrates were

stored at 4 °C in the dark until utilisation to avoid any unexpected biochemical

reactions.

3.2 Experimental methods

3.2.1 Biomethane potential test equipment

The BMP test system constructed for this study consisted of fermentation bottles

submerged in a water bath and a biogas collection gallery (Figure 5 and Figure 6).

The fermentation bottle (Wiltronics Research Pty Ltd) was made of glass and had

volume of 1000 mL. Each bottle was equipped with a rubber bung and an S-shape

air-lock. The air-lock is filled with water to allow biogas to escape but prevent air

from entering the fermentation bottle (Figure 7). The bottle was submerged in the

water bath (Model SWB20D, Ratek Instrument Pty Ltd) to maintain the temperature

at a constant value. Each bottle was connected to a plastic valve and a gas collector

through flexible plastic tube. The biogas collector comprised a 1000 mL plastic

36

measuring cylinder and a plastic container containing NaOH solution for biogas

wash [20, 133, 134] at concentration of 1M in this study. To measure the volume of

CH4 generated from the fermentation bottle, the cylinder was first filled with 1M

[134] NaOH solution, and was inverted and then partially submerged into the NaOH

container. Biogas from the fermentation bottle was introduced into the submerged

part of the cylinder, thus, allowing the NaOH solution to absorb CO2 and H2S from

the biogas. The remaining CH4 gas displaced the NaOH solution inside cylinder and

the CH4 gas volume generated each day can be recorded. The water bath could hold

up to eight fermentation bottles. Up to eight CH4 measuring cylinders could also be

held by the biogas collection gallery. Thus, the constructed BMP test unit could

simultaneously hold eight fermentation bottles. Two identical systems were

constructed and used in this study.

3.2.2 Experimental protocols

To minimize oxygen contamination, prior to the BMP experiment, all fermentation

bottles were continuously flushed with N2 for 5 minutes before filling with 750 mL

of substrates and inoculum. The headspace in each bottle was 250 mL. The bottle

was flushed again with N2 and immediately sealed with the rubber stopper (Figure 7).

After placing into the shaking water bath, which was set at 35.0 ± 0.1 °C, the valve

was opened to allow biogas to enter the gas collection gallery. Shaking velocity of

the water bath was 70 – 75 strokes per minute. The experiment was terminated when

less than 5 mL of CH4 was produced over a day.

37

Figure 5. A picture of the BMP test system.

Figure 6. The schematic diagram of BMP test system.

-

-

-

-

Inverted

measuring

cylinder

Water bath

Bung

Fermentation

bottle

S-shape

air-lock

Plastic basin contains 1M NaOH

38

Figure 7. A picture of one BMP bottle unit used in this study.

3.2.3 Experimental description

The anaerobic batch experiments were initialised to perform BMP tests of substrates,

including single digestion of each raw primary sludge and glycerol, and their co-

fermentation under anaerobic condition. The AD system was then used to evaluate

the TrOC removal efficiency of anaerobic sludge treatment. Figure 8 presents the

experimental scheme of this investigation. Further details are given as following.

39

Figure 8. Experimental roadmap.

3.2.3.1 Batch tests on CH4 potential of sewage sludge and glycerol

Experimental protocol of a BMP assay to evaluate the effect of buffer addition on

AD of sewage sludge alone is described in Table 8. The BMP bottles were seeded

with a mixture of digested sludge and raw primary sludge (known as I/S ratio) of 1/9

by volume. Because of their low pH and alkalinity, 6.2 and 1228.6 mg/L

respectively, sodium bicarbonate (NaHCO3) was added into these bottles to gradually

enhance the buffering capacity at the beginning of AD process. NaHCO3 was added

at increasing concentrations, 15 and 30 mM. All these batch tests were conducted in

duplicate. Data were expressed as mean values.

Table 8. Experimental description of BMP bottles with increasing buffering

capacity.

No Labels of BMP

bottles I/S (by volume) Added NaHCO3 (mM)

1 R-0 1/9 0

2 R-15 1/9 15

3 R-30 1/9 30

Another bottle with only sludge mixtures was made as a baseline for comparison. A

blank bottle with digested sludge was prepared for CH4 yield from digested sludge

only to determine the net CH4 yield. During the experiment, CH4 volume was

recorded daily until CH4 production was detected at less than 5 mLCH4/d.

Anaerobic batch digestion set-up

Biomethane production of

substrates

Anaerobic batch digestion tests

Elimination of TrOCs

AD of three kinds

of glycerol

Anaerobic co-digestion of raw

primary sludge and glycerol

AD of raw

primary sludge

40

Another set-up was used to assess BMP of raw primary sludge under anaerobic

condition. The ratio between inoculum and substrate was increased up to 1/1 by

volume. Moreover, digested sludge used in this set-up was already degassed for the

best microbial activity, namely exhausted digested sludge. There were two BMP

bottles of this mixture but no blank sample as a result. These analyses were

conducted in duplicate. Reading of CH4 production was taken daily.

With the purpose of enhancing the CH4 yield, co-digestion was taken place with pure

glycerol (chemical grade) as a co-substrate for sewage sludge. Batch tests were also

conducted to examine how glycerol seeded bottles perform and stabilise. In terms of

performance, CH4 yield and CH4 production rate were evaluated. Mixture sludge at

ratio of I/S (1/9 by volume) was fed to all bottles in this experiment. Once again,

buffer supplement by NaHCO3 (15 and 30 mM) was applied. Unlike the previous

BMP assays, increasing percentages of pure glycerol (0.5% and 1.0% of raw primary

sludge) were added (Table 9). This addition did not change the whole working

volume. Other two reference bottles without pure glycerol were taken into account

for a baseline of CH4 generation with an introduction of 15 and 30 mM NaHCO3,

while CH4 yield of digested sludge only was given from the previous batch test.

Therefore, three sets of bottles, namely R-15 and R-30 for buffer supplement at 15

and 30 mM without added glycerol, R-15-0.5 and R-30-0.5 for 0.5% glycerol

treatments, and R-15-1.0 and R-30-1.0 for 1.0% glycerol treatments, were

established.

Table 9. Experimental description of BMP bottles with glycerol addition.

No Labels of BMP

bottles

I/S

(by volume)

Glycerol

(%of RS)

Added NaHCO3

(mM)

1 R-0 1/9 0 0

2 R-15 1/9 0 15

3 R-30 1/9 0 30

4 R-15-0.5 1/9 0.5 15

5 R-15-1.0 1/9 1.0 15

6 R-30-0.5 1/9 0.5 30

7 R-30-1.0 1/9 1.0 30

One more series of BMP bottles was set up for identifying the actual CH4 yield of

glycerol. In this case, three different kinds of glycerol, i.e. pure, biodiesel and BIA

glycerol, were utilized as a carbon source for anaerobic microorganisms or digested

sludge. They were determined their viscosity by a rheometer (Section 3.3.5). Their

41

relative high viscosity (at 35 °C) required a well mixing step when glycerol was

introduced into bottles. All BMP bottles were fed with fresh anaerobic inoculum and

then glycerol with increasing percentages, i.e. 0.25% and 0.5% of digested sludge by

volume. Content and labels of BMP bottles are shown in Table 10. Two control

bottles contained only anaerobic inoculum. CH4 volume was recorded daily until

CH4 was produced at rate less than 5 mLCH4/d.

Table 10. BMP bottles assessing BMP of different types of glycerol.

No Labels of

BMP bottles

DS

(mL)

Types of

glycerol

Glycerol

(% of DS by volume)

Glycerol

(mL)

1 P-0.25 750 Pure glycerol 0.25 1.88

2 P-0.5 750 Pure glycerol 0.50 3.75

3 BID-0.25 750 Biodiesel 0.25 1.88

4 BID-0.5 750 Biodiesel 0.50 3.75

5 BIA-0.25 750 BIA 0.25 1.88

6 BIA-0.5 750 BIA 0.50 3.75

Digested sludge (DS) or inoculum

To assess the stabilisation of the AD process, certain parameters such as pH, total

alkalinity, total organic acids, CODt and CODs were measured at the start and end of

each batch test.

3.2.3.2 Anaerobic treatment test of trace organics

The last BMP set-up was for evaluating the fate of trace organics during the

anaerobic treatment of sludge. A set of 30 TrOCs was selected as representatives of

emerging organic groups, namely pharmaceuticals, steroid hormones, phytoestrogens,

UV-filters and pesticides, which are omnipresent in sewage sludge from WWTPs.

The selection was also made according to public concerns regarding their

applications in the agricultural sector and their diversity in terms of chemical

structures and properties (i.e. hydrophobicity and solubility). The hydrophobicity was

identified via their log D values at pH 8.0, including hydrophilic (log D<3.2) and

hydrophobic (log D>3.2) compounds (Table 26 – Appendix). All these TrOCs were

purchased at analytical grade. Their combined stock solution (25 μg/mL) was freshly

prepared in pure methanol. This solution was kept at –18 °C in the dark and used

within one month.

Up to ten BMP bottles used in this arrangement were firstly seeded with anaerobic

inoculum. The same volume of the stock solution of 30 TrOCs was spiked into all

these bottles to achieve a concentration of 135 μg/L of each trace organic

42

contaminant. A gentle stirring step was followed in 5 minutes to obtain a

homogenous spike. This concentration is equal to 5μg/gTS, as TS of anaerobic

inoculum was reported at 27 g/L in this batch test. This investigation was operated for

35 digestion days. At particular time intervals, 0, 2, 6, 9, 14, 21, 28, and 35 days, each

bottle was withdrawn for solid phase extraction (SPE) within 24 hours, and

subsequently TrOC analysis. In this series, two non-spiked bottles, containing only

inoculum, were integrated as blank samples. Stability of this anaerobic mineralization

of TrOCs were governed in terms of parameters, including CH4, TS, VS, COD, pH,

alkalinity, and ammonia nitrogen. All analyses were performed in duplicate.

3.3 Analytical methods

3.3.1 Basic parameters

pH was measured by using a Metrohm Advanced pH/Ion Meter at the room

temperature, around 25 °C. TS, VS, and alkalinity were measured according to the

Standard Methods [135].

For measuring TS and VS concentrations, sludge sample (20 mL) was transferred

into pre-weighted crucibles. The crucibles were then placed in a water bath for 1 hour

at 100 °C for dewatering before being placed in an oven for 24 hours at 105 °C. The

crucible was allowed to cool down to room temperature; and the total weight of the

crucibles and dried sample was measured. Accordingly, TS was determined as the

dried sludge retained in the crucibles. The crucible was then placed into an oven

(1400 furnace, Barnstead Thermolyne, Australia) preheated at 550 °C for 15 minutes.

The crucible was allowed to cool down to room temperature and weighted again to

determine VS concentration.

Supernatant collected from 50 mL sample after centrifugation was used for alkalinity

analysis using the titration method [135]. The solution H2SO4 of0.1N was used as the

titrating chemical and was calibrated using 0.05N Na2CO3. Titration was performed

to a pH end-point of 4.5. Alkalinity was calculated according to the following

equation:

mL sample

, N A /L) (mgCaCOAlkalinity

000503

(1)

where: A = mL of standard 0.1 N H2SO4 used, and

N = normality of standard 0.1N H2SO4

43

Total organic acids were measured by the method 5560C with distillation step

followed by titration step with standard 0.1N NaOH to the end-point pH of 8.3 [135].

Result, expressed by mg acetic acids/L, was from the following equation:

fmL sample

, N mL NaOH /L) cetic acidcids (mg aVolatile a

00060 (2)

where: N = normality of standard 0.1N NaOH, and

f = recovery factor for a given distillation apparatus

3.3.2 Chemical oxygen demand

The determination of CODs was performed by centrifuging a volume of sample (20

mL) and filtering through a Milipore HA membrane 0.45 μm pore size [136].

Approximately 10 mL of sample was also used to measure CODt. After appropriate

sample preparation, each sample and one blank (deionized water) were taken for 0.2

mL and introduced into High Range Plus COD Digestion Reagent vials (200 – 15000

mg/L), and heated at 150 °C in a Hach DBR200 COD Reactor for 2 hours. After this

digestion step, vials were cooled down to room temperature to proceed to

colorimetric determination by using a Hach DR/2000 spectrophotometer with

method 435 COD HR (Figure 9). The result was expressed as mg COD/L.

a) b)

Figure 9. Hach equipment for COD determination: a) DBR200 COD Reactor and

b)DR/2000 Spectrophotometer.

3.3.3 Methane yield calculation

During the experiments, the volume of CH4 generated can be calculated from

measurements of the change in liquid height in the column and container.

44

One blank bottle containing only DS with the equivalent volume of sample bottles

was included to account for the CH4 yield generated from inoculum alone. The net

CH4 volume was corrected by subtracting the CH4 volume in the blank from the CH4

volume in the sample at the same time of sampling:

of blank - CH of sample CHn (mL) productioNet CH 444 (3)

BMP, which is based on the chemical nature of substrates, was determined by

normalising the generated CH4 volume against the initial VS or COD weight basis:

ded (g)D or VS adMass of CO

(mL)tive CHNet cumula yield CHCumulative 4

4 (4)

Specific methanogenic activity (SMA) is calculated as the following [137]:

)./(350

dgVSgCODVSV

RSMA

inoculum (5)

where R is daily CH4 production (mLCH4/d), V and VSinoculum are volume and VS

concentration of inoculum (known as digested sludge), and 350 is a conversion factor

according to Mc Carty [138] that every one gram COD is theoretically converted into

350 mL CH4 in one anaerobic digester.

CH4 yield was modelled using the first order degradation equation according to

[139],

kt

o eGtG 1)( (6)

where G(t) is the cumulative CH4 yield at time t (mLCH4/gVSadded or

mLCH4/gCODadded), Go is the ultimate CH4 yield (mLCH4/gVSadded or

mLCH4/gCODadded), which is defined as the final volume of CH4 beyond which no

more CH4 is released, k is the first order rate constant, and t is the digestion time at

which the corresponding cumulative CH4 production is recorded.

The modified Gompertz equation was established by assuming that the CH4

production in batch condition is a function of growth rate of methanogenic bacteria

over time:

1

max

t)(λ

oG

eR

e

o eGG(t) (7)

where G(t) is the cumulative CH4 yield at time t (mLCH4/gVSadded or

mLCH4/gCODadded), Go is the ultimate CH4 yield (mLCH4/gVSadded or

mLCH4/gCODadded), Rmax is the maximum CH4 production rate (mL/d), λ is the

45

duration of lag phase (d), t is the digestion time at which the corresponding

cumulative CH4 production is recorded.

The fitness of these above equation in regulating the CH4 production was assessed

using the nonlinear curve-fitting tool of Matlab. At the same time, all parameters, the

standard error and the coefficient of determination or correlation coefficient (R2)

were also calculated.

Average values of CH4 yields in all experiments were statistically compared by using

one-way ANOVA (Analysis of Variance) in Excel. The calculated F value and

tabulated F value were compared to judge whether difference between two values is

significant or not. If the F value is greater than F critical or calculated P-value is

greater than level of significance (α), there is at least a significant difference between

two means. In this case, the least significant difference (LSD) of any of them was

calculated at α = 0.05 and α = 0.01 as following equation:

r

stLSD

22 (8)

where tα is tabulated value identified from the degree of freedom for error (df) and α,

s2 is the mean square for error (MS), and r is the number of replications on which the

means were computed. MS and df were calculated by using one-way ANOVA tool in

Excel.

3.3.4 Trace organic analysis

Concentrations of TrOCs in the liquid phase were measured by using the method

reported by Hai et al. [140]. The schematic procedure of these methods is described

in Figure 10.

46

Figure 10. A schematic of sample preparation in solid and liquid phases GC-MS

analysis.

Samples were analysed at the particular times during the experiment. At every single

analysis, standard TrOC solution samples were included for the accurate

measurement. At first, sludge samples were centrifuged to separate supernatant and

sludge pellet. In the supernatant, these compounds were initially extracted from 500

mL sample volume using Oasis HLB cartridges (Waters, Milford, MA, USA). They

were then eluted from the cartridges and the eluents were evaporated to dryness.

After that, the dry residues in the vials were derivatised, cooled to room temperature,

and subjected to gas chromatography–mass spectrometry (GC–MS) analysis using a

Shimadzu GC–MS (QP5000) system. TrOC concentrations in solid phase were

determined according to the method previously described by Wijekoon et al. [141].

An extraction step, including sludge pellet freeze-drying and ultrasonic solvent

extraction was used. Sample was then analysed using the same method for the liquid

phase. All analysis was conducted in duplicate.

3.3.5 Viscosity of glycerol

Viscosities of pure, biodiesel, and BIA glycerol as a function of temperature were

measured by using a Physcia MCR 301 Anton Paar Rheometer (Figure 11). The

Sludge samples

(750 mL)

Centrate Sludge pellet Centrifuge (3500 rpm, 4 o C, 10 min)

Freeze drying for 4h (0.5 g)

Ultrasonic solvent extraction

Decanted supernatant

Remaining sample

Centrifuge at 3500 rpm and 10

min

500 mL aqueous sample Dilute with MQ up to 500 mL

Filter through 0.45 µm filter paper and adjust

pH to 2 – 3

Filtrate (500 mL) for SPE GC – MS

47

rheometer was controlled by the RHEOPLUS/32 V3.40 software. The sample

volume was 1 mL.

Figure 11. A photograph of the Anton Paar Physica MCR 301 Rheometer.

48

4 BIOMETHANE POTENTIALOF SEWAGE SLUDGE AND GLYCEROL

This chapter first describes the key characteristics of sewage sludge and glycerol.

Thereafter, batch-mode BMP tests were performed in order to identify their

biodegradability with respect to CH4 yield. Effects of buffer supplement and

inoculum over substrate (I/S) ratio on CH4 yield and process stability were also

evaluated for sewage sludge, respectively. Meanwhile, the biodegradability of three

types of glycerol, namely pure, biodiesel and BIA, were assessed and compared.

4.1 Sewage sludge

4.1.1 Characteristics of sewage sludge

Key properties of raw primary sludge, digested sludge (which was used as

inoculum), and exhausted digested sludge are summarised in Table 11. The TS and

VS over TS ratio of raw primary sludge were 25.3 g/L (or 2.5%) and 90.4%,

respectively. The TS content of raw primary sludge reported here is consistent with

the literature (for example TS content in the range of 2.0 – 8.0%) as can be seen in

Table 1 of Chapter 2. The raw primary sludge used in this study has a high VS

fraction, which is evidence in a VS/TS ratio that is slightly higher than the range of

60 – 80% typically reported in the literature (Table 1). On the other hand, a lower

VS/TS ratio was found in the digested sludge (ca 60%) compared to raw primary

sludge. This is typical for inoculum. As expected, the digested sludge also had a

much lower soluble organic fraction than the raw primary sludge (i.e. 0.8 gCODs/L

compared to 2.7 gCODs/L).

Values of pH of the raw and digested sludge differ markedly from each other. Due to

the primary treatment process used in the sewage treatment plant, raw primary sludge

was relatively acidic (pH = 5.8) and had a low alkalinity (953.8 mgCaCO3/L). On the

other hand, a high pH and much higher buffering capacity (4185.7 mgCaCO3/L)

were observed in digested sludge. This high alkalinity value reflects the stability of

the AD process. Because of the production of VFAs, which accelerates the

acidification stage, inadequate buffering capacity system may lead to system failure.

In order to sustain the stability of anaerobic conversion process, the ratio of VFAs

and total alkalinity should be maintained at below 0.4 [142]. The VFA/total

alkalinity ratio of the digested sludge reported here was 0.17 (Table 11), indicating

that the inoculum were from a healthy anaerobic digester.

49

Ammonia concentration is another parameter to assess the biological performance of

the AD process. As can be seen from Table 11, raw primary sludge used in this study

had low ammonia content (i.e. 543 mg/L). Regarding the effect of ammonia nitrogen

levels on the anaerobic system, Procházka et al. [143] reported that inherent

buffering capacity inside the anaerobic digester is a function of organic acids,

ammonia and bicarbonate content (Section 2.2.3). While these authors confirmed the

inhibitive effect of high ammonia level of 4 g/L, to microorganism, they also

reported the low CH4 production resulted from 0.5 g/L ammonia nitrogen. These

conclusions were dependent on inoculum origin and types of substrates [58]. In the

present study, the inoculum has a high ammonia content (610 mg/L) which was

possibly produced during the methanogenesis step from nitrogen-containing

substrate [42].

After collection, digested sludge further incubated for 24 days to obtain exhausted

digested sludge (Section 2.2.3). VS/TS ratio of exhausted digested sludge was lower

than that of initial digested sludge. Moreover, its buffer capacity relatively increased

up to approximately 6000 mgCaCO3/L possibly due to the production of ammonia

(i.e. 890 mg/L). This incubation process was considered to be favourably operating

without any inhibition since the VFAs/alkalinity ratio was 0.18. Differences in terms

of pH and COD between initial and exhausted digested sludges were negligible.

Table 11. Characteristics of raw, digested, and exhausted digested sludge.

Characteristics Raw primary

sludge

Digested

sludge

Exhausted digested

sludge

TS (g/L) 25.3 21.3 – 27.0 22.9

VS (g/L) 22.9 13.6 – 17.5 13.4

VS/TS (%) 90.4 63.8 – 64.8 58.5

pH 5.8 7.6 7.6

Alkalinity (mgCaCO3/L) 923.8 4185.7 5947.6

CODt (g/L) 22.7 22.4 21.7

CODs (g/L) 2.7 0.8 0.8

CODs/CODt (%) 11.9 3.6 3.7

NH4+-N (mg/L) 543 610 890

Total organic acids

(mg/L) nd 703.8 1073.1

nd = no data

50

4.1.2 Effects of buffer supplement on the anaerobic treatment of sewage sludge

Batch tests were conducted to assess impacts of buffer on biodegradation of RS

under anaerobic condition (Section 3.2.3). Table 12 lists the average values of key

parameters of raw primary sludge and inoculum mixtures before and after the

experimental period.

An increase in pH and buffer capacity occurred along with increasing concentrations

of NaHCO3 or bicarbonate added (Table 12). The addition of NaHCO3 had no impact

on other parameters, including TS, VS, and COD. A slight increase of pH from 6.2 to

6.6 and 6.9 was a result when supplying 15 and 30 mM NaHCO3 into the initial

sludge mixture, respectively. Correspondingly, the buffer capacities increased from

1228 to 1952 and 2929 mgCaCO3/L. As previously discussed in Section 2.2.3, the

optimal pH range of anaerobic process are from 6 to 8. Meanwhile, alkalinity values

obtained here are in line with those (ca 1000 – 3000 mgCaCO3/L) in the literature

[144]. Therefore, the introduction of NaHCO3 facilitated the bacterial activity at the

start-up of digestion.

Table 12. Characteristics of sludge samples.

Characteristics Sludge mixtures

R-0 R-15 R-30

Conte

nt I/S (by volume) 1/9 1/9 1/9

Added NaHCO3 (mM) 0 15 30

Init

ial

val

ues

TS (g/L) 24.1 27.8 31.5

VS (g/L) 21.3 23.1 25.3

pH 6.2 6.6 6.9

Alkalinity (mgCaCO3/L) 1229 1952 2929

CODt (g/L) 31.4 31.7 29.7

CODs (g/L) 2.0 2.2 2.2

Fin

al v

alu

es TS (g/L) 17.7 17.0 18.3

VS (g/L) 14.0 12.6 12.2

pH 4.4 4.5 4.7

Alkalinity (mgCaCO3/L) 605 930 1442

CODt (g/L) 33.5 31.9 30.6

CODs (g/L) 10.8 12.1 12.6

VS removal (%) 34.1 45.7 51.5

Net CODs (g/L) 8.8 9.9 10.4

Net total alkalinity (mgCaCO3/L) 624 1022 1487

R-0, R-15, and R-30 refer to I/S mixture containing 0, 15, and 30 mM NaHCO3,

respectively

51

The cumulative CH4 generation per unit mass of VS or COD added as a function of

time are presented in Figure 12 and Figure 13, respectively. The obtained data can be

described by both the first order and modified Gompertz models (Figure 14). CH4

production occurred immediately at the beginning of the incubation process, due to

the acclimatisation of inoculum to raw primary sludge (Section3.2.3.1). The

cumulative CH4 production increased until day 6 to 8 and no further CH4 production

could be observed. The short duration of batch test in the present study is also

consistent with that reported by Lim et al. [145]. They concluded that all BMP

bottles ceased working after only 6 to 8 days of biogas production, resulting from

either due to adverse conditions developed or the substrates had been depleted [145].

Because of low VS removal (34.1 – 51.5%) observed after incubation, the effect of

unfavourable conditions, including acidic pH (ca 4) and low buffer capacities (605 –

1440 mgCaCO3/L) on methanogenesis was one of the reasons.

The peaks of CH4 production rates were at the first day of digestion for all BMP

bottles, calculated as 151.7, 296.8, and 291.8 mLCH4/gVSadded.d respectively for R-0,

R-15 and R-30 (Figure 12). This confirms the good adaptation of inoculum to

substrates. The lag times derived from the modified Gompertz model showed the

consistency with the practical observation, varying from zero for R-30 to 0.01 and

0.04 days correspondingly for R-0 and R-15 (Table 13). The data indicated that the

high bicarbonate concentration can shorten the lag phase of methanogenesis, which

was also confirmed by Lin et al. [146] as neutral pH in digester was reached. Hao et

al. [147] investigated the extension of lag phase at pH as low as 5.5. The addition of

NaHCO3, a source of easily degradable substrate, could cause this reduction of lag

phase. Lin et al. [146], however, reported that CH4 yield slow down with increasing

bicarbonate levels, which may be due to used higher-concentrated bicarbonates (150

and 200 mM). CH4 yield from buffered reactors was significantly higher than from

the reactors without bicarbonate addition, which may originate from higher buffer

capacity (15 and 30 mM). In turn, the difference of CH4 yields between bottles with

15 and 30 mM of bicarbonate, was found insignificant (P-value = 0.312 > 0.05,

LSDα=0.05 test1). Comparatively, the ultimate CH4 production did not rely on the level

of bicarbonate although there was a slight increase in COD, demonstrating that low

1Least of significant difference at α = 0.05, was computed using the single-factor ANOVA tool in

Excel 2010

52

NaHCO3 only has a buffering effect. The amount of necessary buffer should be

higher, consequently. In fact, NaHCO3 concentrations (15 and 30 mM) in this study

were still lower than that reported in other studies [146, 148]. Lin et al [146] applied

a range from 0 – 200 mM NaHCO3 with an interval of 50 mM, while 150 mM

NaHCO3 and 150 mM K2CO3 were used by Vavilin et al. [148] from which well-

buffered conditions were maintained. Similarly, in terms of VS basis, R-15 and R-30

bottles had higher performance in methanogenesis with 29.1 and 26.6

mLCH4/gVSadded, respectively, than R-0 bottles with 21.8 mLCH4/gVSadded (Figure

13).

Compared to data reported in certain previous studies (Table 6), these values are

relatively low, which could be attributed to the low concentration of inoculum. A I/S

ratio (or the ratio of digested sludge and raw primary sludge by volume) of 1/9 was

applied in this study. These results are only comparable with those of a study

assessing BMP of thickened sludge from a municipal WWTP by Lim et al. [145].

Using a I/S ratio of 1/8 (by volume), the authors stated that the AD of thickened

sludge had the ultimate CH4 yield of 21.93 mLCH4/gVSadded. They suggested that

methanogenesis was significantly inhibited at low I/S ratio due to high VFA

concentration and an acidic pH.

Figure 14 shows that both first order kinetic and modified Gompertz models can

describe the CH4 yield very well. The higher coefficient R2 (0.990 – 0.996) from the

first order kinetic model in comparison to that (0.970 – 0.991) from the modified

Gompertz model indicated that the best fitting model to the substrates used in this

batch test is the first order kinetic model (Table 13). This model was known to

successfully describe the rate of hydrolysis in the anaerobic degradation [149]. Based

on a direct relationship between the CH4 production and substrate consumption in

anaerobic digesters (Section 2.2.1), this model was also useful to evaluate the kinetic

of this batch test.

53

0 5 10 15 20 25

0

50

100

150

200

0

50

100

150

200

250

300

0 5 10 15 20 25

0

50

100

150

200

CH

4 p

ro

du

cti

on

ra

te (

mL

CH

4/g

VS

ad

ded

.d)

0

50

100

150

200

250

300

Cu

mu

lativ

e C

H4

yie

ld (m

LC

H4

/gV

Sa

dd

ed)

0 5 10 15 20 25

0

50

100

150

200

CH4 production rate Cumulative CH4 yield

Time (d)

0

50

100

150

200

250

300

R1-15

R1-0

R1-30

Figure 12. Cumulative CH4 yield and CH4 production rate for buffered and non-

buffered reactors with increasing buffer concentrations, 0, 15, and 30 mM NaHCO3,

respectively labelled as R-0, R-15, and R-30.

54

0 5 10 15 20 25

0

100

200

300

400

0

10

20

30

0

100

200

300

Cu

mu

lati

ve C

H4 p

ro

du

cti

on

(m

LC

H4

)

Time (d)

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

VS

ad

ded

)

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

CO

Ds)

R1-0 R1-15 R1-30

Figure 13. Temporal variations of cumulative CH4 production of the sample mixture

R1 (1/9 I/S) supplemented with increasing buffer concentrations, 0, 15, and 30 mM

NaHCO3, respectively labelled as R-0, R-15, and R-30.

55

0 5 10 15 20 240

50

100

150

200

250

300

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4

/gC

OD

s)

R-0

R-0 first order model

R-0 modified Gompertz model

R-15



R-30




curves of the first order and modified Gompertz models.

The greatest VS removal, 51.5%, was achieved in R-30 followed by 45.7% in R-0

and 34.1% in R-15. This low percentage of VS reduction suggests that an incomplete

digestion. Meanwhile, all BMP bottles resulted in a high acidic pH (ca 4.5) and low

alkalinity. A digester imbalance occurred due to low pH, which inhibited biomass

activity. This could be referring to the sharp accumulation of VFAs in the start-up of

the digestion course in combination with the insufficient buffer capacity [143, 145,

146]. This observation is in line with that in a study of Procházka et al. [143] who

demonstrated that incomplete substrate destruction with reference to VS removal of

53.6% was observed as a consequence of poor biomass activity, acidic pH, and high

VFA concentration although pH was neutralised by Na2CO3. The authors also

indicated even lower removal of substrate in other reactors without buffering agent,

only from 13.4% to 29.5% VS removal.

Stable performance of anaerobic treatment could be seen with regards of soluble

matter. In particular, initial and final concentrations of CODs and CODt were in the

proportion with increasing amount of supplemented NaHCO3 (Table 12). Whilst no

significant change was observed in CODt concentrations, the values of net CODs

56

(calculated as subtracting the final to initial CODs concentrations) showed an

increasing trend at the end of this batch tests, 8.8, 9.9, and 10.4 mg COD/L

respectively for bottles R-0, R-15, and R-30. Since CODs represents the extent of

solubilisation [84], this tendency could be mainly attributed to the VFA

accumulation since a decrease in total alkalinity (defined as the difference between

final and initial alkalinity) increased proportionally with an increase of

CODsconcentrations (Table 12). This reflects insufficient contributions of acid

consumers compared to acid formers, suggesting that there were a complete

conversion in the hydrolysis/acidification step and a stress situation in the

methanogenic stage. The same conclusion was also stated by Raposo et al. [142,

144].

Table 13. Kinetic analysis of CH4 production in the batch test of buffer

enhancement.

Models and results

(at 95% confidential interval) R-0 R-15 R-30

Experimental cumulative CH4 yield

(mLCH4/gCODs) 221.8 306.0 294.8

First order kinetic model

k (1/d) 0.688 0.896 0.848

R2

0.996 0.995 0.990

Predicted cumulative CH4 yield

(mLCH4/gCODs)

220.2 ±

2.0

302.5 ±

2.9

289.6 ±

3.8

Modified Gompertz model

Lag phase - λ (d) 0.04 0.01 0.00

R2

0.991 0.976 0.970

Maximum CH4 production rate - Rm

(mLCH4/gCODs.d)

97.8 ±

14.0

177.5 ±

45.0

160.8 ±

23.3

Predicted cumulative CH4 yield

(mLCH4/gCODs)

218.3 ±

2.8 300 ± 6.1

286.9 ±

6.4

In terms of the CH4 production and VS degradation, the NaHCO3 supplement

conclusively resulted in a higher performance. Nevertheless, the level of buffering

capacity applied here was still not adequate as can be seen in the short CH4

production duration (Figure 12) and low VS removal efficiency (Table 12).

4.1.3 Methane production of sewage sludge

Initial and final values of parameters of mixture of raw primary sludge and digested

sludge of 1/1 ratio by volume are summarised in Table 14. Exhausted digested

57

sludge used in this batch series was regarded as a healthy anaerobic consortium

rather than a source of substrate. That means CH4 yield was only derived from raw

primary sludge.

Table 14. Characteristics of raw primary sludge and sludge mixture.

Characteristics Raw primary sludge Sludge mixture

Init

ial

val

ues

TS (g/L) 38.9 32.4

VS (g/L) 34.4 25.2

VS/TS (%) 88.6 77.9

pH 5.8 8.1

Alkalinity (mgCaCO3/L) 923.8 3777.0

CODt (g/L) 48.9 35.3

CODs (g/L) 11.0 5.9

CODs/CODt (%) 22.5 16.7

Fin

al v

alues

TS (g/L) nd 16.8

VS (g/L) nd 10.7

VS/TS (%) nd 63.5

pH nd 7.9

Alkalinity (mgCaCO3/L) nd 5329

CODt (g/L) nd 12.5

CODs (g/L) nd 1.9

CODs/CODt (%) nd 15.2

Rem

oval

effi

cien

cies

(%)

TS nd 48.2

VS nd 57.7

CODt nd 64.8

CODs nd 68.8

nd = no data

Slightly lower solid and organic matter content were observed in the raw primary and

digested sludge mixture compared to those in raw primary sludge. The sludge

mixture thus showed lower ratios of VS/TS and CODs/CODt (i.e. 88.6% and 22.5%)

than those in raw primary sludge (i.e. 77.9% and 16.7%). In spite of this reduction,

the usage of 1/1 (v/v) ratio could offer sufficient bacterial activity from inoculum to

well degraded substrate from raw primary sludge. Moreover, the I/S ratio applied in

this batch test lead to an enhancement in buffer capacity in the sludge mixture. In

particular, pH values and alkalinity values in the sludge mixture were observed at 8.1

and 3777 mgCaCO3/L while those values in raw primary sludge were 5.8 and 923.8

mgCaCO3/L, respectively. It was then expected the stable performance of anaerobic

sludge conversion. Data of digestate characteristics also confirmed this process

stability. Neutral pH (7.9) and high alkalinity value (5329 mgCaCO3/L) remained in

58

the expected range for a healthy anaerobic activity. Destruction of both solid and

organic matter was achieved, with the removal efficiencies of TS, VS, CODt, and

CODs were 48.2%, 57.7%, 64.8%, and 68.8%, respectively.

The cumulative CH4 yield of raw primary sludge was presented inFigure 15. A

gradual generation of CH4 lasted 115 days in this batch test, indicating the high

complexity of raw primary sludge. The similar incubation time was applied by [84]

whose quantified the ultimate biodegradability of using the batch-mode test. The

observed pattern of cumulative CH4 production was another indicator for this

complexity. Two separate periods in CH4 production might correspond to more

easily degradable and slowly degradable materials (Figure 15). This was probably

explained by the different biodegradability among organic fractions in raw primary

sludge. This observation revealed the anaerobic inoculum was able to recover after

the suspended time and utilise all organic fractions available in raw primary sludge

subsequently.

0 10 20 30 40 50 60 70 80 90 100 110 120

0

500

1000

1500

2000

2500

3000

3500

4000

4500

5000

5500

6000

Cumulative CH4 production Cumulative CH4 yield

Time (d)

Cu

mu

lati

ve C

H4

pro

du

cti

on

(m

LC

H4

)

0

100

200

300

400

500

600

700

800

900

1000

Cu

mu

lati

ve C

H4

yie

ld (

mL

CH

4/g

VS

ad

ded

)

Figure 15. Profile of methanogenesis of raw primary sludge in terms of production

and yield over experimental time; error bars refer to the standard deviation (n = 2).

Figure 16 shows the fitted curves of two kinetic models, namely the first order and

modified Gompertz models to evaluate experimental CH4 production data and

59

predict cumulative CH4 yields. Comparatively, the regression coefficients from the

first order equation could not be computed, demonstrating its failure in reflecting the

methanogenesis of such complex substrate as raw primary sludge. Meanwhile, by

using the modified Gompertz equation, the CH4 production seemed to be more

accurately described (R2 = 0.987). Moreover, Kim et al. [90] highly recommended to

use another modified Gompertz equation, namely the modified Gompertz model 2.

The authors demonstrated that this model could successfully describe the progress of

CH4 generation of a substrate with diverse organic fractions. In detail, this model was

developed by adding a secondary term reflecting the second lag time of slowly

degradable matter into the modified Gompertz equation. As such, the modified

Gompertz equation was rewritten as following:

1

2

1

1

22

21

1

1

t)(λG

eR

e

t)(λG

eR

e eGeGG(t) (9)

where G(t) is the cumulative CH4 yield at time t (mLCH4/gVSadded); G1 and G2 are

the CH4 production potential in the initial and secondary stages (mLCH4/gVSadded);

R1 and R2 are the maximum CH4 production rate at the initial and secondary stages

(mL/d); λ1 and λ1 are the duration of initial and secondary lag phases (d); t is the

duration of digestion at which the corresponding cumulative CH4 production is

recorded (d).

The curve-fitting plot of these models and corresponding regression parameters are

shown in Figure 16, Figure 17 and Table 15. In a comparison, results obtained from

the modified Gompertz model showed the high goodness of fit (R2 = 0.995). The

estimated lag times derived from this model were 8.4 and 23.1 days for the more

easily degradable and slowly degradable materials, respectively. Apart from a long

suspension in methanogenesis, the low maximum CH4 production rates were

estimated at the initial and secondary stages (i.e. 5.0 ± 0.1 and 12.2 ± 4.8

mLCH4/gVSadded.d). These observations implied that raw primary sludge possessed

organic fractions showing low CH4 production rates. The presence of slowly

biodegradable portion in sewage sludge was confirmed by Luostarinen et al. [82]

who found that the significant CH4 yield was only attained after 10 – 15 days at the

batch mode. The complexity of raw primary sludge used in this study was found. The

different ultimate CH4 yields was calculated from the modified Gompertz model 2

(Table 15).

60

0 20 40 60 80 100 120

0

100

200

300

400

500

600

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4/g

VS

ad

ded

)

Experimental cumulative methane yield

The first order model

The modified Gompertz model


fitting curves of the first order model and modified Gompertz model applied for raw

primary sludge.

0 20 40 60 80 100 120

0

100

200

300

400

500

600

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4

/gV

Sa

dd

ed

)

Experimental cumulative methane yield

The modified Gompertz equation 2


fitting curve of the modified Gompertz model 2 applied for raw primary sludge.

61

Nevertheless, the experimental and predicted ultimate CH4 yield was of high values,

reporting at 594.6 mLCH4/gVSadded (at digestion day 115) and 731.4 ± 38.5

mLCH4/gVSadded (from the modified Gompertz model), respectively. These values

were comparatively higher with a range reported in the literature (Table 6). This

discrepancy was probably due to various operational infrastructures, biogas

collection systems, and calculation methods applied. It was here highlighted that raw

primary sludge was one of potential substrate for CH4 yield.

Stability in methanogenesis of raw primary sludge was observed during this batch

test due to favourable anaerobic conditions (i.e. neutral pH and high buffer capacity).

Experimental results and kinetic analysis data here gave evidences for the complexity

of raw primary sludge in terms of organic fractions with different biodegradability.

Although raw primary sludge were also identified to possess a great potential of CH4

yield, low CH4 production rate and long lag phase made it unfavourable for a well-

established anaerobic digester. As such, the supplement of substrates rich in readily

biodegradable part was prospectively considered to simulate the methanogenesis of

raw primary sludge in particular and sewage sludge in general.

Table 15. Kinetic analysis of CH4 production potential of raw primary sludge.

Models and results (at 95% confidential interval) Raw primary sludge

Experimental cumulative CH4 yield (mLCH4/gVSadded) 594.6


Lag phase - λ (d) 2.412

R2

0.987

Maximum CH4 production rate - Rm

(mLCH4/gVSadded.d) 6.3 ± 0.2

Predicted ultimate CH4 yield (mLCH4/gVSadded) 731.4 ± 38.5

Modified Gompertz model2

Lag phase 1 – λ1 (d) 8.438

Lag phase 2 – λ2 (d) 23.12

R2

0.995

Maximum CH4 production rate – R1


Maximum CH4 production rate – R2


Predicted ultimate CH4 yield – G1

(mLCH4/gVSadded) 890.9 ± 109.8

Predicted ultimate CH4 yield – G2

(mLCH4/gVSadded) 98.5 ± 17.2

62

4.2 Glycerol as a co-substrate

4.2.1 Characteristics of pure and crude glycerol

Three different types of glycerol (namely pure, biodiesel and BIA) were used in this

investigation. Their key properties are summarised in Table 16.

In general, high solubility of glycerol in water made it readily bioaccessible to

microorganisms. The high COD levels (1030 – 1140 g/L) of glycerol, moreover,

indicated its high source of biodegradable carbon, the primary substrate for anaerobic

microbes. These observations are in good agreement with that previously reported

(Section 2.3.1).

Comparatively, crude glycerol investigated in this study had a similar range of

glycerol content with that published in the literature (Table 16). A typical crude

glycerol produced from homogeneous base-catalyzed transesterification possesses 50

– 60% of glycerol (Section 2.3.1). Thompson and He [150] reported that the purity of

crude glycerol from various vegetable oils is from 60 – 70%. In another study, the

purity of crude glycerol from seeds and beans was reported at between 78 – 84%

[151].

Table 16. Physicochemical properties of pure and crude glycerol.

Properties Pure Biodiesel BIA

Appearance Clear Amber Pale brown

Odour Odourless Grain like odour Mild odour

pH 6.7 4 8 – 9

Boiling point (°C) 290 >130 >130

Solubility in water Highly soluble Highly soluble Highly soluble

Na content (mg/L) 0 40 16,939

K content (mg/L) 0 119 454

Specific gravity (at 25 °C) 1.26 1.22–1.24 1.25

Glycerol content (%) ≥ 99 ~75-85 ~50

COD (gO2/L) 1048 1030 1140

BIA = BIA glycerol

Certain key physicochemical characteristics of glycerol can be related to their

impurities. Specific gravity is one instance. It has been determined that the specific

gravity of a 100% pure glycerol is about 1.26 at 25 °C. Of used glycerol, biodiesel

and BIA glycerol had a lower specific gravity (1.22 – 1.25) probably due to the

63

presence some lighter constituents than glycerol, such as hydrocarbons and water

[18, 150]. These impurities might be derived from biodiesel production. At the

operational temperature (35 °C), it was indicated that their viscosity was

proportionally increasing to their purity, with the highest viscosity being observed in

pure glycerol, followed by biodiesel and BIA (Figure 18).

10 15 20 25 30 35 40 45 50 55 60 650.0

0.3

0.6

0.9

0.0

0.3

0.6

0.9

0.0

0.3

0.6

0.9

Temperature (°C)

Pure glycerol

Vis

co

sity

(P

a.s

)

Biodiesel

BIA

10 15 20 25 30 35 40 45 50 55 60 65 700.00

0.01

0.02

0.03

0.04

0.05

0.06

0.07

Vis

co

sity

(P

a.s

)

Temperature (°C)

Figure 18. Viscosity of pure glycerol and two types of crude glycerol (biodiesel and

BIA) as a function of temperature from 10 °C to 60 °C.

Another marginal difference between pure and crude forms of glycerol is a variety of

elements. Sodium and potassium content were here reported at relative greater

64

quantities for BIA glycerol at 16939 and 454 mg/L respectively than for biodiesel

glycerol. They are resulted from catalysts used in processing and neutralisation step

[151]. Applying different biodiesel production processes and purifying approaches,

together with various sources of feedstocks, therefore, brought out a wide range of

purity of this versatile byproduct. Consequently, the presence of these impurities may

lead to the lower CH4 production (Section 2.3.2).

4.2.2 Variations of control parameters in the study of methanogenesis of glycerol

Glycerol was introduced into digested sludge at increasing percentages for evaluating

the kinetic and potential of their CH4 productions. Details of this batch test were

described in Section 3.2.3. Characteristics of this seed (labelled as R0), compositions

and some relevant features of sample mixtures are shown in Table 17. 0.25 and 0.5%

each glycerol were mixed with this fresh inoculum to make up sample mixtures.

They were labelled as P-0.25, P-0.5, BID-0.25, BID-0.5, BIA-0.25, and BIA-0.5

(with P, BID, and BIA denotes pure, biodiesel, and BIA glycerol, respectively).

It should be noted that the introduction of low quantity of glycerol into digested

sludge had a negligible impact on all parameters, except for COD and total organic

acids. Irrespectively of glycerol type, pH varied between 8.1 and 8.7. Likewise, there

was a slight fluctuation among buffer capacity values of all sample mixtures.

Compared to background values of digested sludge, including pH of 7.6 and

alkalinity of 4186 mgCaCO3/L, these obtained values were not significantly

different. In this set-up, all BMP bottles were operated with favourite start-up

conditions, including close to neutral pH and reasonably high alkalinity (ca from

3000 to 4000 mgCaCO3/L). As such, it was expected that methanogenesis lasts

longer to accomplish the ultimate CH4yields.

Both TS and VS content in the sample mixtures experienced a slight increase as

related with glycerol loads. In detail, background concentrations of TS and VS were

21.3 g/L and 13.6 g/L, lower than those of sample mixtures of 23.1 – 25.9 g/L and

15.2 – 17.3 g/L, correspondingly. Likewise, initial values of CODs were proportional

to increasing glycerol concentrations. The addition of 0.25% and 0.5% of glycerol

considerably enhanced the organic matter concentrations, around four times in CODs

concentration (Table 17). The ratio between CODs and CODt, so-called COD yields,

expressing the extent of solubilisation of samples. In the sample mixtures, this

fraction was in the range of 3.8 – 18.5%, greater than that of digested sludge (3.5%).

65

Table 17. Key properties of sample mixtures of three types of glycerol and sludge.

Characteristics Sludge samples

Inoculum (DS) P-0.25 P-0.5 BID-0.25 BID-0.5 BIA-0.25 BIA-0.5

Types of glycerol - Pure Pure Biodiesel Biodiesel BIA BIA

Glycerol (% of DS) - 0.25 0.5 0.25 0.5 0.25 0.5

Init

ial

val

ues

TS (g/L) 21.3 23.1 25.9 24.4 24.5 23.7 24.3

VS (g/L) 13.6 15.3 17.3 16.1 16.3 15.2 16.0

pH 7.6 8.7 8.6 8.7 8.1 8.7 8.7

Alkalinity (mgCaCO3/L) 4186 3521 3401 3928 3378 4331 3950

CODt (g/L) 22.4 23.3 25.8 23.9 26.0 23.0 27.1

CODs (g/L) 0.78 0.9 4.0 1.3 4.8 1.3 1.8

Total organic acids (mg/L) 703.9 1468 1781 1941 1964 2412 2699

CODs/CODt (%) 3.5 3.8 15.5 5.4 18.5 5.7 6.6

NH4+-N (mg/L) 610 nd nd nd nd nd nd

Fin

al v

alues

TS (g/L) 22.9 19.7 21.1 21.5 20.6 19.8 20.7

VS (g/L) 13.4 11.5 13.0 13.2 12.9 12.1 12.7

pH 7.7 8.1 8.1 8.0 8.2 8.2 8.2

Alkalinity (mgCaCO3/L) 5948 5023 5042 5140 5033 5316 5423

CODt (g/L) 21.7 17.0 19.2 22.4 19.5 17.4 16.7

CODs (g/L) 0.75 0.43 0.64 0.58 0.64 0.58 0.68

Total organic acids (mg/L) nd 892 817 799 920 1156 1232

NH4+-N (mg/L) nd 1015 1180 1171 1156

VS removal (%) nd 24.8 24.9 18.0 20.9 20.4 20.6

TS removal (%) nd 14.7 18.1 11.9 15.9 16.5 14.8

CODs removal (%) nd 52.2 84.0 54.0 86.6 54.0 61.8

Digested sludge (DS), nd = no data; all data are expressed as mean values; P-0.25, P-0.5, BID-0.25, BID-0.5, BIA-0.25, and BIS-0.5 refer to

mixtures of DS and pure (P), biodiesel (BID) and BIA glycerol at 0.25% and 0.5% each by volume, respectively

66

Simultaneously, the glycerol introduction increased the organic acid levels. This

observation suggests the availability of organic acid content in three types of

glycerol. As seen in Table 17, greater amount of organic acids was observed in crude

than pure glycerol, referring to higher impurity of crude glycerol. It was pointed out

in some studies that this constituent in residual glycerol could be derived from raw

materials and inhibit bacterial activity [151]. From Figure 19, nonetheless, it can be

seen that anaerobic digesters were expected to be constantly well performed without

any risk of acid accumulation since the ratio of total organic acids and alkalinity was

still within the satisfactory range (0.3 – 0.4).

Given that this anaerobic treatment process was launched with initial adequate

control parameters as discussed above, the stability of the anaerobic conversion was

able to be monitored in some extent by evaluating these parameters at the end. All

final values are presented in Table 17.

The final pH values still laid within the range for normal growth of methanogens (ca

8). Furthermore, relatively higher values of total alkalinity at the end, from 5023 to

5948 mgCaCO3/L. These values may implicate the stable performance of this AD

process. Due to the fact that equilibrium of buffer capacity is created by ammonia

nitrogen, organic acids, and CO2 [143], higher buffer capacities after incubation

could be elucidated by the production of ammonia nitrogen or by the consumption of

organic acids. Higher ammonia nitrogen concentrations at the end, up to around 1000

mg/L, were a consequence of biodegradation of the nitrogen-containing matter.

However, no clear relationship between ammonia nitrogen levels and load added as

well as glycerol types was obtained. Hence, these nitrogen-containing materials were

essentially originated from digested sludge but glycerol, a nitrogen-free substrate

[152]. Additionally, there was no accumulation of organic acids regardless of any

used glycerol, reflecting their effective degradation by anaerobic ecosystem. This

indicated no suppression of methanogenic activity happened and the hydrolytic-

acidogenic stage was successfully completed.

Correspondingly, all bottles showed a clear reduction of CODs. As related to

increasing glycerol concentrations, the CODs removal increased with from around 50

to 80% each kind of glycerol (Table 17). Additionally, removal efficiencies of CODt,

TS, and VS were also achieved at lower extent. All mixtures at different glycerol

types and concentrations showed the similar conversion performance of TS and VS,

11.9 – 18.5% of TS removal and 18.0 – 24.9% of VS removal. The values show

67

anaerobic digesters were healthily functioned at the methanogenesis step to consume

soluble matters from hydrolytic-acidogenic step. Another indicator of sustainable

working of current BMP bottles is the ratio of total organic acids and total alkalinity.

In the current batch test, this ratio ranged between 0.15 and 0.2 at the end of

digestion time, showing well-buffered systems (Figure 19).

P-0.25 P-0.5 BID-0.25 BID-0.5 BIA-0.25 BIA-0.25

1000

1500

2000

2500

3000

3500

4000

4500

5000

5500

P-0.25 P-0.5 BID-0.25 BID-0.5 BIA-0.25 BIA-0.25

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.35

0.40

0.45

0.50

To

tal

alk

an

ity

(m

gC

aC

O3/L

)

Initial values Final values

Th

e r

ati

o o

f to

tal

org

an

ic a

cid

s a

nd

alk

ali

nit

y

Figure 19. Variations of total alkalinity and ratios of total organic acids and total

alkalinity as related to type and concentrations of glycerol; 0.25% and 0.5% addition

of pure (P), biodiesel (BID) and BIA glycerol were labelled as P-0.25, P-0.5, BID-

0.25, BID-0.5, BIA-0.25, and BIA-0.5 respectively; error bars refer to the standard

deviation (n=2).

68

Considering all control parameters, 0.25% and 0.5% of each glycerol applied under

current conditions investigated in the present work are still not beyond the limit

concentration. This is a result of approaching a reasonable I/S ratio, ranging from

1.48 to 1.52 (gCOD/gVS) and from 1.49 to 1.59 (gCOD/gVS) respectively for 0.25%

and 0.5% addition of glycerol.

4.2.3 Methanogenic conversion of glycerol

Figure 20illustrates the cumulative CH4 yield as a function of digestion time with

different types and ratios of glycerol to digested sludge. Data reported here proved

that on average the usage of glycerol, a highly degradable carbon source, stimulated

the anaerobic conversion. CH4 production or substrate utilization rates of all glycerol

treatments mostly showed peaks for the first few days with the highest rate of 0.25%

pure glycerol (193.4 mLCH4/gCOD.d) at the first day. This indicated that anaerobic

microbes were quickly able to digest the glycerol. It could be ascribed to easily

biodegradability of glycerol.

All types of glycerol had the similar fashion of CH4 production, which increased

until around day 10, beyond that a gradual stable trend was reached (Figure 20).

However, two contrasting trends in rates of generating CH4were obtained when

anaerobically treating different types of glycerol. Pure and BIA glycerol showed

higher CH4 potential at lower concentration (0.25%). The time for acclimatisation

could be one of the reasons behind this behaviour. The longer lag phases were

observed in 0.5% compared to that in 0.25% addition of pure and BIA glycerol. This

is due to the fact that bacteria in digested sludge had no previous exposure to

glycerol. On the contrary, higher CH4 production of the digester treating 0.5% than

0.25% biodiesel glycerol was obtained (Figure 20). Residual glycerol from biodiesel

production process had a remarkable methanol content of 8 – 12% by weight [18, 96,

150]. Under anaerobic condition, this constituent is directly converted into CH4.

Accordingly, the more methanol biodiesel glycerol has, the more methanogenic

activity is accelerated.

69

0

100

200

300

400

500

600

700

800

0

50

100

150

200

0

100

200

300

400

500

600

700

800

0

50

100

150

200

0 5 10 15 20 25 30 35 40

0

100

200

300

400

500

600

700

800

0 5 10 15 20 25 30 35 40

0

50

100

150

200

Pure glycerol Pure glycerol

0.25 % each glycerol added 0.5 % each glycerol added

Biodiesel Biodiesel

Net

cu

mu

lati

ve C

H4 y

ield

(m

LC

H4/g

CO

D)

Net

da

ily

CH

4 y

ield

(m

LC

H4/g

CO

D.d

)

BIABIA

Time (d) Time (d)

Figure 20. Cumulative CH4 yield and CH4 production rate of three different types of glycerol, namely pure (P), biodiesel (BID) and BIA, at

increasing concentrations of 0.25% and 0.5%, respectively labelled as P-0.25, P-0.5, BID-0.25, BID-0.5, BIA-0.25, and BIA-0.5.

70

Like the CH4 production rates, similar observations were obtained with cumulative

CH4yield for all types of glycerol. On average, these experimental values were

calculated at 754.5, 449.1, 416.8, 639.5, 720.5, and 303.5 mLCH4/gCOD

respectively for P-0.25, P-0.5, BID-0.25, BID-0.5, BIA-0.25, and BIA-0.5, at day 35.

Data here revealed the comparatively higher performance of BMP bottles to the

values published. Siles López et al. [133] reported cumulative CH4 yields of pre-

treated glycerol achieved the highest point at the lowest loading. They stated that the

acidified glycerol treated with granular sludge exhibited the highest CH4 yield of 323

mLCH4/gCOD. The lowest CH4 yield observed in BIA treatment is in line with these

values. Meanwhile, there is a good agreement between the CH4 yield of 0.25% pure

glycerol and that value reported by Amon et al. [153].

As expected, the higher purity of glycerol, the greater CH4 production of BMP

bottles. The highest CH4 potential (754.5 mLCH4/gCOD) was reached with the pure

form of glycerol even at lower concentration, meanwhile the treatment of 0.5% BIA

resulted in the lowest performance in methanogenesis (303.5 mLCH4/gCOD). There

was, therefore, a correlation of CH4 yield with the impurity of glycerol. Among

substrates, BIA had the lowest pure glycerol and the highest quantity of sodium (16.9

gNa/L), negatively affecting the methanogenesis of this glycerol. Inhibition

concentrations of sodium have been reported widely in the literature [138, 154].

Under anaerobic condition, biomass activity and metabolism of anaerobic microbes

could be readily interfered and then inhibited by high sodium concentration. In the

biological sense, this phenomenon could be explained by the participation of Na+ in

channels on cell membrane and some metabolic steps. McCarty et al. [138]

investigated the effect of sodium concentrations on methanogens, indicating that

moderate to strong inhibition resulted from using 3.5 – 5.5 gNa/L and 8 gNa/L,

respectively. For anaerobic granular biomass, Rinzema et al. [154] reported the

decrease of10, 50, and 100% methanogenic activity was caused by sodium at

concentrations of 5, 10 and 14 g/L, respectively, under mesophilic temperature and

neutral pH.

Plots and kinetic coefficients of the first order and modified Gompertz models are

presented in Figure 21, Figure 22, and Table 18. The soundness of kinetic models

depends on the nature of compounds. It was stated that the first order kinetic model

was suitable for a very simple substrate [155] while the more complex substrate

71

mixture was successfully monitored by using the modified Gompertz model [90].

The complexity of feedstocks according to Kim et al. [90] was due to the mixing of

food waste and sewage sludge. To obtain higher correlation efficiencies, the authors

tested a modified Gompertz model. In this study, data confirmed the fitness of both

kinetic models to CH4 generation of pure and biodiesel glycerol with relative high

correlation (>91%). Results of these kinetic studies showed the short time taken for

initializing the CH4 generation. Regarding BIA, the first order kinetic model is better

fitting to the trend of CH4 production of 0.25% BIA. Alternatively, the modified

Gompertz model showed higher goodness of fit in predicting the methanogenesis of

0.5% BIA. It could be due to this model could not reflect the lag phase of batch tests

properly [90]. Longer lag phase in BMP bottles treating 0.5% BIA was probably

caused by its higher concentration of impurities compared to that of bottles with

0.25% BIA.

0 5 10 15 20 25 30 350

100

200

300

400

500

600

700

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4

/gC

OD

)

P-0.25

P-0.25 first order kinetic model

P-0.25 modified Gompertz model

P-0.5

P-0.5 first order kinetic model

P-0.5 modified Gompertz model


curves of the first order and modified Gompertz models of pure glycerol at two

concentrations of 0.25% (P-0.25) and 0.5% (P-0.5).

72

a)

0 5 10 15 20 25 30 35

0

100

200

300

400

500

600

Time (d)

Cu

mu

lati

ve m

eth

an

e y

iled

(m

LC

H4

/gC

OD

)

BID-0.25

BID-0.25 first order kinetic model

BID-0.25 modified Gompertz model

BID-0.5

BID-0.5 first order kinetic model

BID-0.5 modified Gompertz model

b)

0 5 10 15 20 25 30 350

100

200

300

400

500

600

700

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4/g

CO

D)

BIA-0.25

BIA-0.25 first order kinetic model

BIA-0.25 modified Gompertz model

BIA-0.5

BIA-0.5 first order kinetic model

BIA-0.5 modified Gompertz model


curves using the first order and modified Gompertz models of a) biodiesel (BID) and

b) BIA glycerol at 0.25% and 0.5% concentrations, respectively labelled as BID-

0.25, BID-0.5, BIA-0.25 and BIA-0.5.

73

Theoretical CH4 yield of pure glycerol used in this experiment is calculated at

approximately 428 mLCH4/gCOD using Buswell equation below [156]:

OHCOCHOHHC 224353 5.025.175.1)( (10)

by taking the specific gravity and COD values of pure glycerol of 1.26 and 1,148

gCOD/L, respectively, into account. Data obtained from experimental and predicted

kinetics showed an excess of CH4 yields compared to the theoretical value in case of

0.25% pure glycerol applied (i.e. more than 700 mLCH4/gCOD). Similar

observations were reported in a study on anaerobic co-digesting of glycerol with

sewage sludge [19]. The authors attributed this greater experimental CH4 yield to

active biomass. It was clear that the growth of anaerobic microbes, particularly

methanogens, were simulated by such readily degradable substrates as glycerol.

Meanwhile, the proximate in values of theoretical and practical CH4 yields when

treating 0.5% pure glycerol was found.

Table 18. Kinetic analysis of CH4 production during the batch test of BMP of

glycerol.

Models and kinetic

parameters

(at 95% confidential

intervals)

Pure glycerol Biodiesel

glycerol BIA glycerol

0.25% 0.5% 0.25% 0.5% 0.25% 0.5%

Experimental cumulative

CH4 yield

(mLCH4/gCOD)

754.5 449.1 416.8 639.5 720.5 303.5


First order rate constant

– k (1/d) 0.410 0.163 0.261 0.200 0.189 0.131

R2

0.982 0.934 0.941 0.983 0.921 0.915

Predicted cumulative CH4

yield – Go

(mLCH4/gCOD)

738.5 ±

8.7

458.2

± 18.6

390.4 ±

10.0

640.6

± 11.0

668.4 ±

24.6

314.2

± 19.6


Lag phase – λ (d) 0.154 1.558 0.000 0.185 0.000 3.057

R2

0.985 0.976 0.913 0.988 0.879 0.993

Maximum CH4

production rate – Rm

(mLCH4/gCOD.d)

208.3 ±

26.4

66.1 ±

9.0

72.5 ±

10.7

86.4 ±

8.2

93.6 ±

15.4

57.0 ±

6.3

Predicted cumulative CH4

yield – Go

(mLCH4/gCOD)

732.1 ±

7.8

442.6

± 7.7

381.4 ±

11.2

627.0

± 7.8

642.2 ±

24.9

292.9

± 3.6

On basis of data of control parameters as related to glycerol types and concentrations

suggests, the introduction of 0.25% and 0.5% of three types of glycerol still

74

demonstrated the adequate function of both hydrolytic-acidogenetic and

methanogenetic bacteria. On the other hand, the profile of both experimental and

predicted CH4 yields implicated an obvious dependence of CH4 yields of glycerol on

its degree of impurities (i.e. sodium concentration). It could be concluded that

glycerol could be considered as a feasible feedstock for the AD process owning to

their high CH4 potential. They might be potentially used as a high-yield co-substrate

for enhancing the anaerobic treatment of low-carbon substrates.

4.3 Summary

In this batch test, the potential of CH4 yields and process stability of the anaerobic

conversion of each sewage sludge and glycerol was tested with different buffer

concentrations and I/S ratios.

In terms of system stability, data from experimental observations indicated

methanogens was favoured by the buffer (NaHCO3) supplement. Compared to the

non-buffed bottles, higher performance in producing CH4 was recorded in bottles

with addition of 15 and 30 mM NaHCO3. Lag times estimated from the first-order

and modified Gompertz models gave evidence of such stimulation of CH4 conversion

by buffer addition. Buffer concentrations applied in this study was, however, still not

adequate for stable methanogenesis, possibly due to low inoculum. An increase of

I/S ratio from 1/9 to 1/1 led to the stability in the methanogenesis of raw primary

sludge due to favourable anaerobic conditions (i.e. neutral pH and high alkalinity),

implying an important role of I/S ratio. In this steady state, results of CH4 yield from

experimental data and kinetic analysis demonstrated a great CH4potential of raw

primary sludge. However, low CH4 production rate observed here suggested a need

of readily biodegradable substrate in order to accelerate the methanogenesis of raw

primary sludge in particular and sewage sludge in general.

Another series of BMP tests of glycerol, which has been stated a potential substrate

for anaerobic digestion, was carried out with the adequate inoculum

supplementation. Evaluation of control parameters among three types of glycerol at

different concentrations showed a stable function of anaerobic conversion. It could

be expected from their high experimental and predicted CH4 yields that glycerol

could be a feasible co-substrate for sewage sludge.

75

5 ANAEROBIC CO-DIGESTION OF SEWAGE SLUDGE AND GLYCEROL

A series of anaerobic co-digestion batch tests was performed using raw primary

sludge and glycerol as a co-substrate. It is hypothesized that the addition of glycerol

increases the carbon content of the feed mixtures, resulting in higher CH4 yield. The

effects of glycerol and buffer (NaHCO3) addition on system stability were

simultaneously assessed.

5.1 Characteristics of mixtures of sewage sludge and glycerol

The feeds, which was the mixtures of raw primary sludge and glycerol with the

supplement of buffer (NaHCO3) were characterised for their key characteristics in

Table 19. This table also presents relevant properties of the seed, or digested sludge,

and other sludge samples of raw primary sludge with and without supplied NaHCO3

for comparison. Experimental procedure in this batch test was stated in Section

3.2.3.1. The mixtures of DS and RS at ratio of 1/9 (R-0) was supplemented with 15

mM and 30 mM NaHCO3, one of which was then introduced with 0.5% and 1.0%

glycerol, respectively, labelled as R-15-0.5, R-15-1.0, R-30-0.5, and R-30-1.0.

Reference bottles containing increasing buffer concentrations and no glycerol were

also utilised for co-digestion evaluation. They were R-0, R-15 and R-30,

characteristics of which were discussed in Section 4.1.2.

Similar changes in TS and VS content were observed in BMP bottles supplemented

with NaHCO3 and glycerol. A slightly increase of VS and TS content in glycerol-

added BMP mixtures (i.e. 0.5% and 1.0% pyre glycerol) with buffer compared to

those adding only buffer (i.e. 15 and 30 mM NaHCO3). Consequently, when VS/TS

ratios were evaluated, no variation among these mixtures was observed, between

80.4% and 88.4% (Table 19).

There was no clear effect of the supplement of glycerol on both pH and alkalinity.

Changes of pH of buffered bottles with glycerol added and that of buffered bottles

including no glycerol was negligible irrespectively of glycerol concentrations (Table

19). At 15 mM NaHCO3 added, pH were 6.6 without glycerol addition and changed

to 6.5 after introducing 0.5% and 1.0% of glycerol. Correspondingly, these values

were 6.9 and 6.8 in case of 30 mM NaHCO3. Compared to pH variation, alkalinity

values experienced the same fashion (Figure 23), demonstrating only NaHCO3

proved its effect on enhancing buffer capacities of co-digestion mixtures as expected.

76

Table 19. Physical and biochemical characteristics of sample mixtures in the anaerobic co-digestion test.

Characteristics Sample mixtures

Inoculum (DS) R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

Conte

nt I/S (by volume) nd 1/9 1/9 1/9 1/9 1/9 1/9 1/9

Glycerol (% of RS volume) nd Nd nd 0.5 1.0 nd 0.5 1.0

NaHCO3 (mM) nd Nd 15 15 15 30 30 30

Init

ial

val

ues

TS (g/L) 21.3 24.1 27.8 34.4 34.0 31.5 34.0 31.9

VS (g/L) 13.6 21.3 23.1 29.5 29.0 25.3 27.9 26.8

VS/TS (%) 63.8 88.4 83.1 85.8 85.3 80.3 82.1 84.0

pH 7.6 6.2 6.6 6.5 6.5 6.9 6.8 6.9

Alkalinity (mgCaCO3/L) 4186 1229 1952 1959 2090 2929 2709 2490

CODt (g/L) 22.4 31.4 31.7 34.5 37.3 29.7 37.8 39.9

CODs (g/L) 0.78 15.5 2.2 9.7 13.6 2.2 9.1 15.5

CODs/CODt (%) 3.5 6.4 6.9 28.2 24.0 7.4 36.4 38.9

Fin

al v

alues

TS (g/L) 22.9 17.7 17.0 27.5 25.9 18.3 24.2 26.8

VS (g/L) 13.4 14.0 12.6 22.4 20.8 12.2 17.8 21.1

VS/TS (%) 58.5 79.1 74.1 81.5 80.3 66.7 73.6 78.7

pH 7.7 4.4 4.5 4.4 4.4 4.7 4.6 4.9

Alkalinity (mgCaCO3/L) 5948 605 930 837 930 1442 1116 1758

CODt (g/L) 21.7 33.5 31.9 45.9 34.3 30.6 30.7 35.3

CODs (g/L) 0.75 10.8 12.1 13.7 14.0 12.6 15.0 16.4

CODs/CODt (%) 3.5 32.2 37.9 29.8 40.8 41.2 48.9 46.2

VS removal (%) nd 34.1 45.7 24.0 28.3 51.5 34.5 21.0

Digested sludge (DS); raw primary sludge (RS); nd = no data, all data are expressed as mean values; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0

refer to the sludge mixture (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5% and 1.0% glycerol each,

respectively

77

R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

0

1

2

3

4

5

6

7

pH

Initial pH Final pH Initial alkalinity Final alkalinity

500

1000

1500

2000

2500

3000

3500

4000

Alk

ali

nit

y (

mg

Ca

CO

3/L

)

Figure 23. Values of pH and alkalinity as related to the glycerol and buffer addition

before and after experimental time; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer to

the mixture of 10% digested sludge and 90% raw primary sludge by volume (R-0)

supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5%

and 1.0% glycerol each, respectively.

As discussed previously (Section4.1.2), 15 and 30 mM NaHCO3had an important

role in well buffering the AD process by increasing pH and alkalinity levels without

causing any change of the other parameters. As such, the start-up of this co-digestion

test was facilitated with close to neutral pH and relatively high alkalinity due to the

NaHCO3 supplementation.

Apparently, increase of the biodegradable organic content for BMP bottles

weremainly contributed by organic matter of glycerol. Whilst the glycerol

supplementation caused a slight increase of TS, VS, and CODt, the liquid part of all

co-digestion mixtures had a considerably greater CODs compared to the mono-

digestion bottles (Table 19). Figure 24 shows thatCODs concentration increased

proportional to glycerol addition (i.e. 0.5 and 1.0% v/v glycerol added) at any buffer

concentrations (i.e. 15 and 30 mM NaHCO3). Soluble organic fraction (CODs) values

of mixtures were around four times and seven times higher respectively after adding

0.5 and 1.0% glycerol than mixtures without glycerol addition. Accordingly, there

78

was a corresponding increase in the ratio of CODs over CODt concentrations with the

addition of glycerol, indicating an improved extent of solubilisation in anaerobic co-

digesters. This observation was expected due to the high solubility and purity of

glycerol. There was a good agreement between data obtained here and that widely

reported elsewhere. Several studies have noticed that the introduction of glycerol had

a positive impact on the organic matter in sample mixtures[19, 20, 53, 152, 153, 157-

159]. According to Wohlgemut et al. [157], the COD load was gradually improved

by introducing increasing percentage of both pure and crude glycerol (i.e. 0.5, 1, 2

and 4% by weight) into reactors treating hog manure. A steady increase in COD level

was also stated by Astals et al. [53] when using crude glycerol as a co-fermentation

additive of pig manure. Siles et al. [20] carried out an increase COD load from 185 ±

5 gCODs/L to 300 ± 5 gCODs/L in the water phase by adding 15% glycerol (by

volume) to wastewater from biodiesel production. An even eight times higher in

CODs concentration was also observed in co-digestion bottles (3% crude glycerol

added) compared to reference digester [152, 158].

R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

0

5

10

15

20

25

30

VS concentration CODs concentration VS/TS ratio CODs/CODt ratio

Co

ncen

tra

tio

n (

g/L

)

0

10

20

30

40

50

60

70

80

90

100

Ra

tio

(%

)

Figure 24. Changes of VS, CODs, and ratios of VS/TS and CODs/CODt in relation

with glycerol and buffer addition; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer to the

mixture of 10% digested sludge and 90% raw primary sludge by volume (R-0)

79

supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5%

and 1.0% glycerol each, respectively.

Glycerol is highly soluble in water while NaHCO3 addition improves the buffer

capacity. Hence, they were expected to create favourable conditions for anaerobic

conversion in terms of system stability and CH4 production performance.

5.2 Removal of volatile solids and chemical oxygen demand

The characteristics of digestates in this batch test are presented in Table 19. Stability

and performance of single-digestion and co-digestion with glycerol of raw primary

sludge were evaluated via the variations of VS and COD.

In comparison to the feed, after anaerobic metabolism, the digestate of all BMP

bottles have a low pH value in the range of 4.5 – 4.9. This acidic pH range, which

could be interpreted as a build-up of organic acids, can induce the system instability

[142, 160-162]. It is not surprising that buffer capacities were deteriorated in all

digestates (Figure 23). Therefore, NaHCO3 concentrations in this study

unsatisfactorily buffered the AD process. Buffer supplementation, nevertheless, still

exhibited its function since significant differences of alkalinity levels were found

with different NaHCO3 concentrations added to the system. Available alkalinity

values were of 837 – 930 mgCaCO3/L at 15 mM NaHCO3 introduced while bottles

buffered with 30 mM NaHCO3 had alkalinity more than 1000 mgCaCO3/L.

Co-digestion bottles exhibited lower VS content removal than the reference bottles.

Final VS concentrations of BMP bottles with glycerol were significantly higher than

those of co-digestion BMP bottles (Figure 25). At 15 mM added buffer, in particular,

final VS values were 12.6 g/L, 22.4 g/L, and 20.8 g/L as 0%, 0.5%, and 1.0%

addition of glycerol. These observations could be elucidated by two factors. First, the

supplementary organic carbon source from glycerol simulated the biomass growth

related to an increase in VS content. This is also stated by Ma et al. [159] who found

a VS increase of 3 g/L in the UASB reactor with the glycerol supplementation after

33 digestion days. In a more recent study, Fountoulakis et al. [19] reported that an

increase of VS concentration, from 17.9 ± 0.8 g/L to 23.7 ± 0.7 g/L after adding

glycerol was a results of biomass growth stimulation. Second, some particulate

organic matter remained in the digestate of co-digestion bottles could be another

reason. This conclusion was clearly confirmed with data of VS/TS ratios, ranging

from 70% to 80%, which generally were of 60 – 70% in a typical anaerobic digester.

80

Therefore, the hydrolytic bacteria in reference bottles better functioned. Astals et al.

[158] had the same observation and explained by the fact that bacteria in co-digestion

reactors favourably utilised glycerol, while biomass might digest particulate organic

content in raw primary sludge in the absence of glycerol.

The variation of VS removed after digestion in relation to the addition of glycerol

isshown in Figure 25. There was no notable difference of VS removal efficiencies

between buffered and non-buffered BMP bottles. A decreasing trend of VS

destruction was found when BMP bottles were fed with increasing glycerol

concentration. Of BMP bottles buffered with 30 mM NaHCO3, the highest

performance of degrading VS was achieved at 51.5% in the bottle R-15, followed by

34.5% and 21.0% in R-30-0.5 and R-30-1.0 respectively. Likewise, at 15 mM

NaHCO3, bottles with 0%, 0.5%, and 1.0% glycerol added degraded 45.7%, 24.0%,

and 28.3% of available VS correspondingly. The VS elimination was one of

indicators of well hydrolysing particulate organic matter in substrate.

R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

0

5

10

15

20

25

Fin

al

VS

co

ncen

tra

tio

ns

(g/L

)

Final VS concentration VS/TS ratio Removal efficiencies

0

10

20

30

40

50

60

70

80

90

100

VS

/TS

ra

tio

(%

)

0

10

20

30

40

50

60

70

80

90

100

Rem

ov

al

eff

icie

ncie

s (%

)

Figure 25. Profile of solid content, including final VS concentrations, VS/TS ratio,

VS removal efficiency at the end of digestion; R-15-0.5, R-15-1.0, R-30-0.5, R-30-

1.0 refer to the mixture of 10% digested sludge and 90% raw primary sludge by

volume (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and

introduced with 0.5% and 1.0% glycerol each, respectively.

81

In addition, a considerable increase of CODs after digestion process, a representative

of soluble matter, was observed (Figure 26). Accumulation of VFAs, an intermediate

product during anaerobic conversion, was considered as a reason owning to an acidic

pH of digestates. A conjunction of VS destruction and VFA build-up here pointed

out an adequate hydrolysis and acidogenesis, and simultaneously a visible suspension

of methanogens.

When physicochemical features of influents and effluents of co-digestion and

reference BMP bottles are taken into account, a reduction in pH to acidic values (ca 4

– 5), a destruction of buffer capacities (ca 1000 mgCaCO3/L), and a considerable

increase in CODs (generally from VFA accumulation) were observed. Data reported

here suggested an obvious instability of anaerobic system; an inadequate efficiency

with regard of VS degradation was a consequence. Simultaneously, it is possible to

relate this destabilisation to CH4 formation although hydrolysis and acidogenesis

were considered well-functioned.

R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

0

2

4

6

8

10

12

14

16

CO

Ds

co

ncen

tra

tio

n (

g/L

)

CODs concentration CODs/CODt

0

10

20

30

40

50

60

70

80

90

100

CO

Ds/

CO

Dt

(%)

Figure 26. CODs concentrations and CODs/CODt ratios in digestates; R-15-0.5, R-

15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 10% digested sludge and 90% raw

primary sludge by volume (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3

(R-30), and introduced with 0.5% and 1.0% glycerol each, respectively.

82

5.3 Methane potential of co-digestion mixture of sewage sludge and glycerol

The extent and yield of CH4 production were experimentally observed during the

anaerobic co-digestion of raw primary sludge with glycerol in a comparison with

reference bottles. Their estimated results from two kinetic models were also included

in order to predict and evaluate the effects of glycerol on the CH4 conversion of raw

primary sludge.

CH4 production in all BMP bottles reached its peak at the first day of digestion time

(Figure 27). Higher peaks of CH4 production were observed with the co-digestion

bottles compared to the reference bottles (R-0) containing raw primary sludge only.

On average, with15 mM NaHCO3 supplement, the highest CH4 production rates were

reached at 267.5 and 255 mLCH4/d for R-15-0.5 and R-15-1.0, respectively, while

175 mLCH4/d was the highest rate of methanogenesis from R-0. This indicated that

bacteria immediately consumed glycerol. Thus, the addition of glycerol in the feed

accelerated the hydrolysis stage. These observations was probably ascribed to high

particulate matter in raw primary sludge, which are hardly transported into bacteria

cells for further degradation, lengthening CH4 production [56]. After getting the

peaks, CH4 production of co-digestion bottles appeared to significantly slow down

over time. Despite exhibiting the highest CH4 production at the first day (340

mLCH4/d), bottles introduced with 1.0% glycerol essentially ceased forming CH4 in

short digestion time, day 4 and around day 6 for 15 mM and 30 mM NaHCO3 added

bottles, respectively. As discussed previously, there was a clear relation of system

stability and CH4 production. In the digestate of R-30-1.0 bottles, unfavourable

conditions for methanogens, including a drop of pH (from 6.9 to 4.9) and a

destruction of buffer capacity (from 2490 to 1758 mgCaCO3/L) could elucidate the

short methanogenic activity. Correspondingly, these bottles showed the lowest VS

removal efficiency (21%) although they were fed with the highest glycerol (1.0%)

and buffer concentrations (30 mM NaHCO3). Similar tendency in terms of pH drop,

alkalinity decrease, and inadequate VS removal was observed with R-15-1.0 bottles

(Table 19). Meanwhile, bottles treating lower glycerol concentrations had the gradual

trend in generating CH4 and maintained digestion process until around day 14. There

would be hence a shock loading for a sustainable anaerobic metabolism with higher

dosage of glycerol (1.0%).

83

0

20

40

60

80

100

0

20

40

60

80

100

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

VS

rem

ov

al)

R-15 R-15-0.5 R-15-1.0

R-30 R-30-0.5 R-30-1.0

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

CO

Da

dd

ed

)

Time (d)

Time (d)

Figure 27. Profile of daily and cumulative CH4 production in the co-digestion batch test; R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer to the

mixture of 10% digested sludge and 90% raw primary sludge by volume (R-0) supplemented with 15 (R-15) and 30 mM NaHCO3 (R-30), and

introduced with 0.5% and 1.0% glycerol each, respectively.

84

When glycerol content was augmented in co-digestion bottles from 0.5 – 1.0% (v/v)

of raw primary sludge, CH4 production on basis of COD fed was found significantly

higher in reference bottles compared to bottles containing glycerol. Irrespective of

buffer concentrations added, R-15 and R-30 bottles had around six and seven times

higher than corresponding bottles with 0.5% and 1.0% glycerol regarding cumulative

CH4 yields (Figure 28). In spite of an enhanced rapidly biodegradable organic

content in the feed, an incomplete conversion of substrate occurred in co-digestion

bottles, particularly in the methanogenesis. Insufficient buffer capacity for temporary

VFA accumulation was probably one explanation for this system failure. It was even

clear with a considerable increase of CODs in their digestates, which was mainly

contributed by VFAs generated from all bottles, especially glycerol-supplied bottles

(Table 19). The condition was not suitable for methanogenic activity due to the low

I/S ratio as reported normally less than 2 [144, 163]. The early termination of CH4

production and inadequate CH4 yields were a consequence of only 10% by volume of

digested sludge used in this investigation.

On the other hand, higher CH4 volume produced as one gram of VS degraded was

accomplished in bottles after the addition of glycerol in the feed throughout the

whole experimental time. When 30 mM NaHCO3was introduced into BMP bottles,

CH4 yields were 57.9, 70.8, and 95.3 mLCH4/gVSremoval from BMP bottles with 0%,

0.5%, and 1.0% addition of glycerol, respectively (Figure 28). Given the fact that the

CH4 production was proportional to the substrate destruction in a given anaerobic

reactor, data given here suggested that co-digesting raw primary sludge with glycerol

improved the anaerobic biodegradation of solid matter into CH4. Nonetheless, it was

again noted that the CH4 yields and VS removal efficiencies were low compared to

what have been previously reported (Table 6). This observation could explain the

system instability of all co-digestion and reference bottles as previously discussed

(Section 5.2).

85

0

20

40

60

80

100

0

20

40

60

80

100

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

VS

rem

ov

al)

R-15 R-15-0.5 R-15-1.0

R-30 R-30-0.5 R-30-1.0

Cu

mu

lati

ve C

H4

yie

ld

(mL

CH

4/g

CO

Da

dd

ed

)

Time (d)

Time (d)

Figure 28. Profile of cumulative CH4 production on basis of VS removal (mLCH4/gVSremoval) and COD added (mLCH4/gCODadded); R-15-0.5,

R-15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of digested sludge and raw primary sludge (1/9 by volume) supplemented with 15 (R-15) and

30 mM NaHCO3 (R-30), and introduced with 0.5% and 1.0% glycerol each, respectively.

86

The first order and modified Gompertz models were tested for cumulative CH4

production in terms of COD added for anticipating further long-term effects of using

glycerol as a co-substrate of raw primary sludge (Figure 29 and Figure 30). Table 20

summaries regression parameters from the first order and modified Gompertz models

for the CH4 production in all reference and co-digestion bottles. Estimated results

revealed that both kinetic models successfully monitor CH4 yields in all cases with

high correlation factors (>98%) except for bottles supplied with 15 mM NaHCO3 and

0.5% glycerol with lower correlation coefficients (84 – 90%).

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

Time (d)

Cu

mu

lati

ve C

H4

yie

ld (

mL

CH

4/g

CO

D)

R-0



R-15



R-15-0.5

R-15-0.5 first order model

R-15-0.5 modified Gompertz model

R-15-1.0




curves of the first order and modified Gompertz models; R-15-0.5, R-15-1.0 refer to

the mixture of 1/9 I/S supplemented with 15 mM NaHCO3 (R-15) and introduced

with 0.5% and 1.0% glycerol respectively.

The first order rate constants (k) were computed at higher values for bottles treating

1.0% glycerol compared to reference bottles; k values were 1.37 and 0.72 1/d in

bottles R-15-1.0 and R-15 respectively, for example (Table 20). This indicated that

the glycerol addition significantly activated the CH4 generation. In fact, an instable

situation was created in all BMP bottles with early cease of producing biogas, the

anaerobic microbes firstly utilised biodegradable matter. This assumption could

clearly elucidate the high soundness of the first order model.

87

0 2 4 6 8 10 12 14 16 18 20

0

50

100

150

200

250

300

350

Time (d)

Cu

mu

lati

ve C

H4

yie

ld (

mL

CH

4/g

CO

D)

R-0



R-30



R-30-0.5 vs. t



R-30-1.0 vs. t




curves of the first order and modified Gompertz models; R-30-0.5, R-30-1.0 refer to

the mixture of digested sludge and raw primary sludge (1/9 by volume)

supplemented with 30 mM NaHCO3 (R-30) and introduced with 0.5% and 1.0%

glycerol respectively.

In order to more properly evaluate the trend of CH4 formation over experimental

period, the modified Gompertz model was tested. From this model, two other

important parameters, including CH4 production rate and lag phase time, were

obtained. The lag phase time was clearly seen from 1.0 % glycerol-added bottles,

reported at 0.14 and 0.17 days with 15 and 30 mM NaHCO3, respectively; while the

other bottles immediately produced CH4. In co-digestion bottles, ultimate CH4 yields

decreased along with increased glycerol addition (0.5 and 1.0%). By contrast, the

maximum CH4 production rates seemed to be achieved higher with more glycerol

fed. CH4 production was therefore simulated but not well maintained over digestion

time. It is clear that the influence of glycerol addition on the methanogenesis with

respect to yields and stability. In a comparison with reference bottles, higher rates of

forming CH4 were obtained from R-15 and R-30 bottles; however, lower CH4

production rates were found in these buffered bottles with glycerol addition (Table

20). This is because the anaerobic conversion of glycerol led to rapid VFA

88

accumulation, declining buffer capacities of co-digestion bottles. The

methanogenesis was quickly suspended as a result. Therefore, necessary amount of

buffer should be higher for better methanogenic performance in co-digestion context.

Table 20. Kinetic analysis of CH4 production on COD basis in the batch test of co-

digestion of raw primary sludge and glycerol.

Models and kinetic

parameters

(at 95% confidential

intervals)

R-0 R-15 R-15-

0.5

R-15-

1.0 R-30

R-30-

0.5

R-30-

1.0

Experimental cumulative

CH4 yield

(mLCH4/gCOD)

281.7 366.8 60.7 49.7 351.3 58.5 34.5


First order rate

constant – k (1/d) 0.570 0.719 0.528 1.367 0.689 0.422 1.793

R2

0.996 0.986 0.865 0.993 0.983 0.907 0.998

Predicted cumulative

CH4 yield – Go

(mLCH4/gCOD)

275.5

± 3.4

356.3

± 7.1

55.9

± 4.0

48.6

± 0.6

340 ±

7.7

55.0

± 0.5

34.4

± 0.2


Lag phase – λ (d) 0.000 0.000 0.000 0.138 0.000 0.000 0.168

R2

0.983 0.961 0.840 0.988 0.956 0.888 0.996

Maximum CH4

production rate – Rm

(mLCH4/gCOD.d)

97.5 ±

20.4

153.9

± 49.3

8.8 ±

4.8

47.4 ±

15.3

138.4

± 46.0

8.2 ±

3.6

42.2 ±

11.4

Predicted cumulative

CH4 yield – Go

(mLCH4/gCOD)

271.8

± 6.6

352.3

± 11.7

57.8

± 5.4

48.4

± 0.8

336.5

± 12.0

56.1

± 4.7

34.4

± 0.3

R-15-0.5, R-15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 1/9 I/S supplemented

with 15 (R-15) and 30 mM NaHCO3 (R-30), and introduced with 0.5% and 1.0%

glycerol each, respectively

For the better understanding of the whole AD process in this investigation, SMA

should be taken into account. SMA was a useful tool to evaluate the anaerobic

population capacity to convert substrate into CH4 in given conditions [134, 137, 151,

164]. In the early stage, the buffer expressed its role in simulating bacteria activity;

with the higher SMA values were achieved in bottles buffered with NaHCO3 than

bottles without buffer (R-0) (Figure 31). BMP bottles with 30 mM NaHCO3 achieved

the highest SMA values of 0.91 and 0.96 gCOD/gVS.d, meanwhile the methanogenic

activity of 15 mM NaHCO3-added bottles was calculated at 0.75 and 0.72

gCOD/gVS.d for 0.5% and 1.0% treatments, respectively. There was an

89

independence between glycerol dosage and methanogenic activity, glycerol

accordingly caused no inhibitory effects on methanogenic consortium.

R-0 R-15 R-15-0.5 R-15-1.0 R-30 R-30-0.5 R-30-1.0

0.0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

SM

A (

gC

OD

/gV

S.d

) The first day The first four days From day 5 onward

Figure 31. Profile of SMA of reference bottles and co-digestion bottles; R-15-0.5, R-

15-1.0, R-30-0.5, R-30-1.0 refer to the mixture of 1/9 I/S supplemented with 15 (R-

15) and 30 mM NaHCO3 (R-30), and introduced with 0.5% and 1.0% glycerol each,

respectively.

It has been reported that higher SMA values could result in higher the rates of CH4

generation on VS basis [137]. There was a clear correspondence between of SMA

and CH4 production rate as far as their patterns over the experimental time were

concerned. The maximum SMA values were at the first day for all co-digestion and

reference bottles, followed by a gradual decrease in bottles without glycerol but an

immediate drop in glycerol-contained feeds. Furthermore, when the SMA values in

this investigation were averaged for the first five digestion days, healthier biomass

activity was observed in bottles containing no glycerol (R-0, R-15, and R-30). From

day 5 onwards, on average, all BMP bottles remained low methanogenic activity;

even a suspension of CH4-forming bacteria was found in R-30-1.0 bottles (Figure

31). In the long-term operation, it can be concluded that this reduction in SMA was

associated to excess organic matter and the insufficient bacterial function, expressed

90

as the low I/S ratio of 1/9 by volume tested. This finding was confirmed by Souto et

al. [134] who found that SMA values were more dependent on the ratio of food and

microorganism than substrate concentration only. Moreover, considering SMA

greatly depends on the particular operational conditions and digestion duration [134,

145], it is evitable to further investigate various ratios of inoculum amount and

glycerol concentrations under definite conditions to determine the threshold for a

stable AD process.

In summary, the supplement of 0.5% and 1.0% of glycerol into raw primary sludge

was satisfactory in terms of CH4 production within the start-up stage. This

improvement resulted from the enhancement of organic source with regard to CODs,

the highly biodegradability and solubility of glycerol. Nonetheless, certain

characteristics of digestates, typical of higher CODs, acidic pH and inadequate

alkalinity as well as short methanogenic duration, indicated the incomplete anaerobic

biodegradation for the prolonged operation. Therefore, in terms of system stability,

additional carbon matter from glycerol only proved its positive effects on the

hydrolytic-acidogenetic stage rather than methanogenesis. These results indicated

that glycerol could be used as a co-substrate; however, the appropriate I/S ratio and

buffer supplementation should be further studied in order to prevent overloading and

attain the ultimate CH4 potential.

5.4 Summary

Data from this anaerobic co-digestion test of raw primary sludge and glycerol

pointed out that the addition of 0.5% and 1.0% of glycerol into raw primary sludge

was satisfactory in terms of daily and cumulative CH4 production within the start-up

stage. This improvement resulted from the enhancement of organic source with

regard to CODs, the highly biodegradability and solubility of glycerol. On the other

hand, certain characteristics of digestates, namely higher CODs, acidic pH and

inadequate alkalinity as well as short methanogenic duration, indicated the

incomplete anaerobic biodegradation for the long-term operation. These findings

indicated effects of I/S ratio and buffer supplementation on system stability and then

the ultimate CH4 potential. Their appropriate levels should be further investigated to

optimize CH4 generation of anaerobic co-digestion.

91

6 FATE OF TRACE ORGANIC COMPOUNDS DURING ANAEROBIC

DIGESTION OF SEWAGE SLUDGE

This chapter aims to evaluate the removal efficiency of TrOCs under anaerobic

condition and their biodegradation as well as distribution between the aqueous

(water) and solid (sludge) phases. The overall performance of the AD process was

examined with respect to biogas production and COD removal, followed by the

removal of these TrOCs and their behaviour in the water and solid phases.

6.1 Overall performance of AD of trace organics

The overall performance of the anaerobic degradation of trace organic-spiked sludge

was evaluated by focusing on the evolution of the variation of relevant parameters,

namely pH, alkalinity, ammonia nitrogen, organic acids, VS/TS ratio, VS and CODs,

and the production of CH4 throughout the digestion time.

6.1.1 The variation of control parameters during over the digestion time

Table 21 presents key characteristics of digested sludge and digested sludge spiked

with TrOCs. Data from BMP experiments using only digested sludge were used as

the reference. This table also shows the variations of these characteristics in TrOC-

spiked bottles at a particular time during the digestion time.

After spiking, no considerable change in most parameters was observed. In

particular, pH values were 7.8 for reference and TrOC-spiked bottles; accordingly,

alkalinity values in theses bottles were similar (4233 and 4316 mgCaCO3/L for

digested sludge without and with TrOCs, respectively). TS and VS content were

slightly higher in reference bottles compared to spiked bottles (27.0 gTS/L and 17.5

gVS/L in bottles before spiking and 25.8 gTS/L and 16.3 gVS/L after spiking,

respectively). However, the VS/TS ratio remained unchanged and was 64% in both

cases. CODs increased significantly with the addition of TrOCs, due to the addition

of the stock solution of TrOCs. The amount of TrOCs (135 μg/L) used here did not

cause any change in CODs concentration. The increase in CODs concentration is

from methanol, which was used as the solvent for the preparation of TrOC stock

solution.

92

Table 21. Characteristics of digested sludge bottles and TrOC-spiked bottles

throughout the experimental time.

Characteristics Sample mixtures

DS R-0 R-2 R-6 R-9 R-14 R-21 R-28 R-35

Conte

nt

Anaerobic

digested

sludge (mL)

750 750 750 750 750 750 750 750 750

TrOCs (μg/L) nd 135 135 135 135 135 135 135 135

Day of

sampling

(nth

day)

nd 0 2 6 9 14 21 28 35

Par

amet

ers

TS (g/L) 27.0 25.8 25.6 26.4 24.3 24.8 24.1 23.7 24.2

VS (g/L) 17.5 16.3 15.9 16.5 15.3 15.0 14.6 14.1 14.5

VS/TS (%) 64.9 63.3 62.1 62.2 62.9 60.4 60.5 59.7 59.9

pH 7.8 7.9 7.6 7.7 7.8 7.8 7.8 8.3 7.8

Alkalinity

(mgCaCO3/L) 4233 4316 4211 4389 4237 4115 4665 2374 2656

CODt (g/L) 28.4 26.0 29.6 28.2 21.3 23.1 20.7 21.7 19.5

CODs (g/L) 0.44 3.51 3.61 1.67 0.89 0.95 1.07 0.76 1.08

CODs/CODt

(%) 1.5 13.5 12.2 5.9 4.2 4.1 5.2 3.5 5.5

Digested sludge (DS); nd = no data; R-0, R-2, R-6, R-9, R-14, R-21, R-28, and R-35

refer to TrOC-spiked digested sludge bottles sampled at day 0, 2, 6, 9, 14, 21, 28,

and 35, respectively; All data are expressed as mean values

Throughout the experiment, values of chemical control parameters as a function of

time were in the desirable range for the AD process. The pH varied in the range of

7.6 – 8.3, which was favourable for the normal growth of anaerobic microbes. The

stable values of pH can be attributed the relatively high buffer capacities of TrOC-

spiked bottles with an initial alkalinity values of 4316 mgCaCO3/L. Figure 32

presents the variation of pH and alkalinity during the course of digestion. Alkalinity

values remained stable at the value of more than 4000 mgCaCO3/L until day 21 and

reduced to more than 2000 mgCaCO3/L for the remaining period of digestion. These

values are compatible with those previously reported for the normal growth of

anaerobic microbes [44].

93

Dig

este

d slu

dge

Day

0

Day

2

Day

6

Day

9

Day

14

Day

21

Day

28

Day

35

7.4

7.6

7.8

8.0

8.2

8.4

8.6

pH Alkalinity

pH

Digestion time

2000

2500

3000

3500

4000

4500

5000

Alk

ali

nit

y (

mg

Ca

CO

3/L

)

Figure 32. Profile of pH and alkalinity values in digested sludge bottles and spiked

digested sludge bottles at the particular digestion time.

Because digested sludge was used as the inoculum, the reductions in TS and VS over

35 days were small (Table 21). However, the removal of soluble organic matter

measured by CODs was significant (Figure 33). The solubilisation and acid

formation of methanol in the hydrolytic-acidogenic stage were responsible for such

increase. From day 2, a significant decrease in CODs could be observed reaching a

stable concentration at around 1gCODs/L after day 9. Due to an insignificant

variation of CODt over the digestion time, it is not surprising that the ratio of CODs

and CODt exhibited the same pattern with CODs during the course of incubation.

94

Dig

este

d slu

dge

Day

0

Day

2

Day

6

Day

9

Day

14

Day

21

Day

28

Day

35

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

CODs CODs/CODt

CO

Ds (

g/L

)

Digestion time

0

2

4

6

8

10

12

14

16

18

20

CO

Ds/C

OD

t (%

)

Figure 33. Profile of CODs and ratio of CODs/CODt of digested sludge and digested

sludge spiked with TrOCs from ay 0 to day 35 of digestion.

6.1.2 Anaerobic conversion of trace organic compounds

Figure 34 shows the daily and cumulative CH4 produced from tested bottles versus

those from reference bottles. Throughout the experiment, the similar performance of

methanogenesis in terms of daily production was observed in both reference and

spiked bottles. The only exception is that CH4 production from bottles containing

TrOCs was significantly higher to that of the reference bottles from day 4 to 9. This

initial high CH4 yield in TrOCs-spiked bottles could be attributed to the addition of

methanol, which was used as the solvent in preparing the TrOC stock solution.

Methanol is a readily amenable carbon source. After the biodegradation of methanol

in day 5 to day 8, which resulted in a high CH4 production rate, the process returned

to normal, and CH4 productions from the TrOC spiked and unspiked bottles were

similar. These results confirm that the addition of TrOCs in methanol did not cause

any inhibitory effects to the AD process.

95

0

500

1000

1500

2000

2500

3000

0 5 10 15 20 25 30 35

0

100

200

300

400

500

600

700

Cu

mu

lati

ve m

eth

an

e p

ro

du

cti

on

(m

LC

H4)

Digested sludge Digested sludge with TrOCs

Da

ily

meth

an

e p

ro

du

cti

on

(m

LC

H4/d

)

Time (d)

Figure 34. Profile of daily and cumulative CH4 production of reference and TrOC-

spiked MBP bottles over 35 days; error bars refer to the standard deviation (n = 2).

The first order and modified Gomperzt kinetic models were used to describe the CH4

yields as a function of time (Figure 35). The first order kinetic model could not

96

closely match the trend of CH4 yields from TrOC-spiked bottles (that contained

methanol) but could successfully describe the CH4 yields of the reference bottles.

This is possibly due to the fact that the microbial consortium in digested sludge used

in this batch test was not previously acclimatised with methanol, resulting in a lag

time of CH4 production. The inclusion of lag time parameter in the modified

Gompertz allows it to describe very well the CH4 yields in all cases in this study

(Table 22).

0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

160

180

200

Time (d)

Cu

mu

lati

ve m

eth

an

e y

ield

(m

LC

H4/g

CO

Ds)

TrOC-sipked digested sludge

The first order equation for spiked sludge

The modified Gompertz equation for spiked sludge

Methanol

The first order equation for methanol

The modified Gompertz equation for methanol

Digested sludge

The frst order equation for digested sludge

The modified Gompertz equation for digested sludge

Figure 35. Plots of cumulative CH4 yields on VS basis and regression fitting curve

using the first order and modified Gompertz models applied for digested sludge,

TrOC-spiked digested sludge, and methanol.

Based on the modified Gompertz equation, the lag time were 0, 1.6, and 4.9 days for

digested sludge, TrOC-spiked digested sludge, and methanol, respectively. In

comparison, the longer lag time for the methanogenesis of methanol suggested the

inhibitory effects of methanol on anaerobic microbes although methanol has been

known as a readily degradable substrate for methanogens [165]. The CH4 production

from bottles treating TrOCs was only suspended for around the first 1.6 days.

Meanwhile, it is clear that the rapid CH4 generation was found in reference bottles, in

which digested sludge already reached its stable performance. Considering the

97

process performance in terms of parameter variations and CH4 yields, a steady state

of methanogenesis was created in TrOC-spiked bottles from day 1.6 onward.

Table 22. Kinetic analysis of CH4 production on VS basis in the batch test of the AD

treating trace organic contaminants.

Models and kinetic parameters

(at 95% confidential intervals)

TrOC-spiked

digested sludge

Digested

sludge Methanol

Experimental cumulative CH4 yield at

day 35 (mLCH4/gVSadded) 216.3 112.0 104.4


First order rate constant – k (1/d) 0.125 0.092 0.146

Correlation coefficient – R2

0.946 0.976 0.803

Predicted cumulative CH4 yield –

Go (mLCH4/gVSadded) 218.6 ± 10.5 109.5 ± 4.4

114.4 ±

10.7


Lag phase – λ (d) 1.591 0.000 4.897

Correlation coefficient – R2

0.974 0.933 0.996

Maximum CH4 production rate –

Rm (mLCH4/gVSadded.d) 24.7 ± 3.9 6.8 ± 0.6 57.0 ± 7.8

Predicted cumulative CH4 yield –

Go (mLCH4/gVSadded) 205.1 ± 4.9 101.8 ± 4.4 57.8 ± 5.4

6.2 Dynamics of trace organics under anaerobic sludge treatment

The digested sludge used in this study was obtained from a full-scale plant. Thus,

TrOCs were detected in the water and solid phases of digested sludge prior to spiking

(Figure 36). However, these background concentrations were small. With the

exception of carbamazepine and triclosan, their concentrations in the solid phase

were well below 5 µg/g.

In wastewater sludge, TrOCs could adsorb onto the solid phase (sludge) and remain

in the water phase [166]. Therefore, both water and solid contributions made up the

concentration of TrOCs in sludge samples. Concentrations of TrOCs in all sludge

samples are calculated as:

CwCsC (11)

and TSXCs (12)

where X (µg/g) and Cs (µg/L) refer to TrOC concentrations in solid phase; Cw

(µg/L) are the concentrations of TrOCs in liquid phase, and TS is the solid

concentration of samples (g/L).

98

0

5

10

15

20

25

30

35

Co

ncen

tra

tio

n i

n t

he w

ate

r p

ha

se (

µg

/L) Water phase

Clo

fibri

c ac

idS

alic

yli

c ac

idK

etopro

fen

Fen

opro

pN

apro

xen

Met

ronid

azole

Ibupro

fen

Pri

mid

one

Dic

lofe

nac

Gem

fibro

zil

Pro

poxur

Form

ononet

inE

nte

rola

ctone

Car

bam

azap

ine

Pen

tach

loro

phen

ol

Est

rone

Atr

azin

eA

met

ryn

Ben

zophen

one

Am

itri

pty

line

4-t

ert-

buty

lphen

ol

Oxyben

zone

Est

riol

Bis

phen

ol

A17-a

-et

hin

yle

stra

dio

l17-b

-est

radio

lT

ricl

osa

n17-b

-est

radio

l-17-a

ceta

te4-t

ert-

oct

ylp

hen

ol

Oct

ocr

yle

ne

0

1

2

3

4

5

6

7

8

9

10

Co

ncen

tra

tio

ns

in t

he s

oli

d p

ha

se (

µg

/g) Solid phase

Figure 36. Distribution of the selected TrOCs in the water and solid phases of

digested sludge.

During digestion, mass transfer of TrOCs between the aqueous and solid phases as

well as biodegradation regulated the evolution of their concentrations. Their mass

99

transfer into air phase (known as volatilisation) was neglected due to their Henry’s

law constant of tested compounds were calculated to be lower than 1.0 × 10-

3atm.m

3/mol (Table 26 – Appendix). The abiotic loss by photodegradation of TrOCs

in this study was also minimised by covering BMP bottles to avoid light exposure.

The mass transfer of TrOCs then virtually occurred between water and solid

compartments of sludge. Hence, the fate of these compounds during the AD of

sludge, two-compartment matrix, involves their elimination by biodegradation and

the mass transfer from the water phase to the solid phase (sorption) and vice versa

(desorption). Their dynamic mode could be schematically presented in Figure 37.

Figure 37. The schematic diagram of the fate of TrOCs in a sludge-water system.

The two-phase fate model can be simplified based on the following assumptions:

Anaerobic biomass growth is negligible along with the steady operation of all

BMP bottles; accordingly, biomass (expressed as VS content) could be

considered stable over experimental time; and

There is no inhibition to methanogenic populations caused by TrOCs as

discussed above due to the process stability over experimental time.

As a result, the behaviour of each compound in this study was only examined via

anaerobic biodegradation and mass transfer in and between two phases of sludges.

When stability of biomass growth was stable, the pseudo first-order kinetic was

preferred to describe biodegradation rate from the water phase only as following

[167]:

CVSkr bb (13)

where rb is the biodegradation rate (μg/L.d), kb is the biodegradation rate constant in

water (L/gVS.d), VS is biomass concentration (gVS/L), C is the initial concentration

of TrOC (μg/L).

Biodegradation (kb)

Solid (sludge) phase

Aqueous (water) phase

Sorption (ks) Desorption (kd)

Water-solid

partition

(kp)

100

The transfer rate between the water phase and solid phase followed the linear

isotherm. The sorption rate was calculated as:

CwTSkr ss (14)

where rs is the sorption rate (μg/gTS.d), ks is the sorption rate constant (L/gTS.d); TS

is total solid (gTS/L); and Cw is the initial concentration of TrOC in the water

phase(μg/L); meanwhile, desorption rate was:

Cskr dd (15)

where rs is the desorption rate (μg/L.d), kd is the desorption rate constant (1/d); Cs is

the initial concentration of TrOC in the solid phase (μg/L).

As an instantaneous equilibrium was created, there was an equal amount of TrOCs

transferring between two phases at an infinite time:

TSCw

Cs

k

kk

CskCwTSk

rr

d

sp

ds

ds

(16)

where kp is the water-solid partition coefficient (L/g).

The dynamic of each TrOC concentration with time was expressed as a differential

equation, in the water phase:

CVSkCsk

kCwTSk

dt

dCw

CVSkCskCwTSkt

C

dt

dCw

b

p

ss

bds

(17)

and in the solid phase:

Csk

kCwTSk

dt

dCs

CskCwTSkt

C

dt

dCs

p

ss

ds

(18)

The two-phase fate model associated to the dynamic of each TrOC compound in two

phases was expressed as a system of differential equations:

Csk

kCwTSk

dt

dCs

CVSkCsk

kCwTSk

dt

dCw

p

ss

b

p

ss

(19)

101

Optimisation of the model parameters in accordance with the experimental data was

carried out by using least-squares regression in Matlab.

The total concentration of one specific compound over time was the sum of

concentrations in the water and solid phases, expressed as:

CVSkdt

CsCwdb

)( (20)

The variation of TrOC concentrations over the digestion time followed the pseudo

first-order equation:

tVSk

otbeCC

(21)

whereCt and Co is the initial concentration (t = 0) and the remaining substrate

concentration at time t (µg/L), t is the time period (d), and kb is the biodegradation

rate constant (L/gVS.d), VS is biomass concentration (g/L).Co was the sum of the

available concentration in water and solid phases after TrOCs were spiked into

samples at the time t = 0:

ooo CwCsC (22)

Removal efficiency (R%) at time t was calculated as:

%100

o

to

C

CCR (23)

Their half-life time, defined as a time that 50% concentration of TrOCs remained, is

calculated as:

VSkt

b

2ln2/1 (24)

where t1/2 is half-life time (d).

Table 23 summarises value of ks, kp, kb, and t1/2 determined from the two phase fate

model using experimental data presented in Figure 38.

102

Table 23. Calculated model parameters achieved from the two-phase fate model.

Group Compound Log D at

pH = 8

Model parameters

ks kp kb t1/2

Phar

mac

euti

cals

Salicylic acid –1.14 0.169 0.011 0.0229 2.0

Ketoprofen –0.55 0.138 0.009 0.0009 51.2

Naproxen –0.18 0.175 0.016 0.1243 0.4

Metronidazole –0.14 0.080 0.116 0.0003 137.4

Ibuprofen 0.14 0.131 0.038 0.0000 -

Primidone 0.83 0.154 0.010 0.0009 50.0

Diclofenac 1.06 0.091 0.075 0.0042 11.1

Gemfibrozil 1.18 0.142 0.029 0.0003 162.8

Carbamazepine 1.89 0.092 0.102 0.0013 36.7

Amitriptyline 3.21 0.076 0.119 0.0000 -

Triclosan 4.92 0.078 0.148 0.0027 16.9

Ste

roid

horm

ones

Estriol 2.53 0.089 0.110 0.0005 84.0

Estrone 3.62 0.112 0.076 0.0018 25.1

17-α-ethinylestradiol 4.11 0.086 0.113 0.0005 92.4

17-β-estradiol 4.15 0.087 0.097 0.0034 13.4

17-β-estradiol-17-

acetate 5.11 0.129 0.056 0.0539 0.9

Pes

tici

des

Clofibric acid –1.29 0.129 0.039 0.0003 140.8

Fenoprop –0.28 0.124 0.021 0.0013 35.8

Propoxur 1.54 0.120 0.035 0.0026 18.1

Pentachlorophenol 2.19 0.065 0.140 0.0171 2.7

Atrazine 2.64 0.113 0.031 0.0078 5.9

Ametryn 2.97 0.078 0.115 0.0036 12.9

Indust

rial

chem

ical

s 4-tert-butylphenol 3.39 0.089 0.109 0.0000 -

Bisphenol A 3.64 0.088 0.112 0.0000 -

4-tert-octylphenol 5.18 0.083 0.117 0.0000 -

Phyto

estr

ogen

Enterolactone 1.88 0.130 0.016 0.0100 4.6

Formononetin 1.81 0.109 0.052 0.0202 2.3

UV

filt

ers Benzophenone 3.21 0.088 0.112 0.0000 -

Oxybenzone 3.42 0.156 0.058 0.1797 0.3

Octocrylene 6.89 0.081 0.118 0.0000 -

Sorption rate constant (ks, μg/L.d), water-solid partition coefficient (kp, L/g),

biodegradation rate constant (kb, L/gVS.d), half-life time (t1/2, d); values of log D at

pH = 8 were sourced from Scifinder Scholar database (ACD/Labs)

103

0 5 10 15 20 25 30 350

50

100

150

200

250

Time (d)

Conce

ntr

atio

n (

µg/L

)

Triclosan

Fitted curve for water phase concentration

Water phase concentration

Fitted curve for sludge phase concentration

Sludge phase concentration

0 5 10 15 20 25 30 35

0

10

20

30

40

50

60

70

80

90

100

Time (d)C

once

ntr

atio

n (

µg/L

)

Pentachlorophenol





0 5 10 15 20 25 30 35

0

50

100

150

200

250

300

350

Time (d)

Conce

ntr

atio

n (

µg/L

)

Metronidazole





0 5 10 15 20 25 30 350

20

40

60

80

100

120

140

160

180

Time (d)

Conce

ntr

atio

n (

µg/L

)

Amitriptyline





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

Time (d)

Conce

ntr

atio

n (

µg/L

)

Octocrylene





0 5 10 15 20 25 30 35

0

50

100

150

Time (d)

Conce

ntr

atio

n (

µg/L

)

4-tert-octylphenol





104

0 5 10 15 20 25 30 350

50

100

150

200

250

300

350

400

450

Time (d)

Conce

ntr

atio

n (

µg/L

)

Ametryn





0 5 10 15 20 25 30 35

10

20

30

40

50

60

70

80

90

100

Time (d)C

once

ntr

atio

n (

µg/L

)

4-tert-butylphenol





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

Time (d)

Conce

ntr

atio

n (

µg/L

)

17-a-ethinylestradiol





0 5 10 15 20 25 30 350

50

100

150

Time (d)

Co

nce

ntr

atio

n (

µg/L

)

Bisphenol A





0 5 10 15 20 25 30 35

0

50

100

150

Time (d)

Co

nce

ntr

atio

n (

µg/L

)

Benzophenone





0 5 10 15 20 25 30 35

20

30

40

50

60

70

80

90

100

Time (d)

Con

centr

atio

n (

µg

/L)

Estriol





105

0 5 10 15 20 25 30 3520

40

60

80

100

120

140

160

180

200

220

Time (d)

Conce

ntr

atio

n (

µg/L

)

Carbamazapine





0 5 10 15 20 25 30 35

0

2

4

6

8

10

12

14

16

Time (d)C

once

ntr

atio

n (

µg/L

)

17-b-estradiol





0 5 10 15 20 25 30 35

0

50

100

150

200

250

Time (d)

Conce

ntr

atio

n (

µg/L

)

Estrone





0 5 10 15 20 25 30 350

20

40

60

80

100

120

140

160

180

Time (d)

Con

centr

atio

n (

µg

/L)

Diclofenac





0 5 10 15 20 25 30 35

-5

0

5

10

15

20

25

30

Time (d)

Co

nce

ntr

atio

n (

µg

/L)

Oxybenzone





0 5 10 15 20 25 30 35

0

2

4

6

8

10

12

14

16

18

20

Time (d)

Co

nce

ntr

atio

n (

µg

/L)

17-b-estradiol-17-acetate





106

0 5 10 15 20 25 30 350

20

40

60

80

100

120

Time (d)

Conce

ntr

atio

n (

µg/L

)

Enterolactone





0 5 10 15 20 25 30 35

0

25

50

75

100

125

150

Time (d)C

once

ntr

atio

n (

µg/L

)

Clofibric acid





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

Time (d)

Conce

ntr

atio

n (

µg/L

)

Ibuprofen





0 5 10 15 20 25 30 350

10

20

30

40

50

60

70

80

Time (d)

Co

nce

ntr

atio

n (

µg

/L)

Propoxur





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

Time (d)

Co

nce

ntr

atio

n (

µg

/L)

Atrazine





0 5 10 15 20 25 30 35

20

30

40

50

60

70

80

90

100

110

Time (d)

Co

nce

ntr

atio

n (

µg/L

)

Gemfibrozil





107

0 5 10 15 20 25 30 350

20

40

60

80

100

120

140

Time (d)

Conce

ntr

atio

n (

µg/L

)

Fenoprop





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

160

180

Time (d)C

once

ntr

atio

n (

µg/L

)

Naproxen





0 5 10 15 20 25 30 35

0

10

20

30

40

50

60

70

Time (d)

Conce

ntr

atio

n (

µg/L

)

Formononetin





0 5 10 15 20 25 30 350

50

100

150

Time (d)

Conce

ntr

atio

n (

µg/L

)

Salicylic acid





0 5 10 15 20 25 30 35

0

20

40

60

80

100

120

140

160

Time (d)

Conce

ntr

atio

n (

µg/L

)

Ketoprofen





0 5 10 15 20 25 30 35

0

10

20

30

40

50

60

70

Time (d)

Conce

ntr

atio

n (

µg/L

)

Primidone





Figure 38. Plots of experimental data and fitting curvesof the two-phase fate model for the selected TrOCs.

108

6.3 Removal performance of trace organic compounds under anaerobic

sludge treatment

6.3.1 Overall removal of trace organic compounds

Biodegradability under anaerobic condition of the TrOCs investigated here

represented by kb varies significantly (Table 24). Some TrOCs (i.e. primidone,

amitriptyline, 4-tert-butylphenol, bisphenol A, and 4-tert-octylphenol) are complete

recalcitrant to anaerobic digestion (kb ≈ 0 L/gVS.d) whilst several others (i.e.

formononetin, pentachlorophenol, enterolactone, salicylic acid, 17-β-estradiol-17-

acetate, naproxen, and oxybenzone) show significant biodegradability (kb> 0.01

L/gVS.d). The other TrOCs exhibit moderate removal capacity.

TrOC removal was determined after 35 days of anaerobic digestion and compared to

literature data obtained from both anaerobic and aerobic conditions (Table 24). Like

observations of biodegradation rates, above99% removal was achieved with all

phytoestrogens (i.e. formononetin, enterolactone), pentachlorophenol, salycilic acid,

17-β-estradiol-17-acetate, naproxen, and oxybenzone while industrial chemicals

showed no observable removal efficiency. For the other classes of TrOCs, the

removal efficiency varied over a wide range. Results reported in this study were

mostly comparable with data recorded in previous studies under anaerobic condition.

On the other hand, the anaerobic digestion is less effective for removing TrOCs

when compared to data from aerobic treatment processes reported in the literature

(Table 24).

Similar removal efficiency from this investigation and previous studies could be

observed with some pharmaceuticals and steroids. Of these TrOCs, however,

carbamazepine and diclofenac are exceptions. They have been reported either to be

resistant to biological removal under aerobic or anoxic conditions [25, 141, 168, 169]

or to be moderately removed only when the seed sludge was long acclimatised [25].

This study, by contrast, shows 48.3% and 88.8% removal of carbamazepine and

diclofenac, respectively, after 35 days of anaerobic digestion. The results here

suggest that carbamazepine and diclofenac are persistent to aerobic treatment but are

amendable to anaerobic biodegradation.

109

Table 24. Overall removal efficiencies of the selected TrOCs in comparison with reported data under anaerobic and aerobic conditions.

Category Compound

This study Literature

Anaerobic Aerobic

kb

(L/gVS.d) t1/2 (d) R (%) after 35 d R (%) References R (%) References

Phar

mac

euti

cals

Salicylic acid 0.0229 2.0 > 99.0 >99.0 [170] 89.0 [141]

Ketoprofen 0.0009 51.2 37.7 15.0 [168] 70.5 – 89.0 [141, 169]

Naproxen 0.1243 0.4 > 99.0 70.0 – 99.0 [25, 168,

171] 40.1 – 78.0 [141, 169]

Metronidazole 0.0003 137.4 16.2 not available 68.5 [141]

Primidone 0.0009 50.0 38.5 < 0.1 – 2.0 [168, 172] 12.4 – 95.0 [141, 169]

Ibuprofen 0.0000 - < 0.1 <0.1 [168] 96.7 – 99.0 [141, 169]

Diclofenac 0.0042 11.1 88.8 <0.1 – 60.0 [24, 25, 168,

171] 17.3 – 27.0 [141, 169]

Gemfibrozil 0.0003 162.8 13.8 not available 91.0 – 99.0 [141, 169]

Carbamazepine 0.0013 36.7 48.3 <0.1 – 5.0 [24, 25, 168,

172, 173] < 0.1 – 58.0 [141, 169]

Amitriptyline 0.0000 - <0.1 47.0 [168] 78.0 – 97.8 [141, 169]

Triclosan 0.0027 16.9 76.2 90.0 [168] 44.0 – 91.8 [141, 169]

Ste

roid

horm

ones

Estriol 0.0005 84.0 25.1 <0.1 [168] 84.0 – 98.2 [141, 169]

Estrone 0.0018 25.1 62.0 <0.1 [168] 97.0 – 98.0 [141, 169]

17-α-

ethinylestradiol 0.0005 92.4 23.1 15.0 [168] 84.0 – 93.5 [141, 169]

17-β-estradiol 0.0034 13.4 83.5 60.0 [168] 99.0 – 99.4 [141, 169]

17-β-estradiol-17-

acetate 0.0539 0.9 > 99.0 not available 98.0 [141]

110

Category Compound

This study Literature

Anaerobic Aerobic

kb

(L/gVS.d) t1/2 (d) R (%) after 35 d R (%) References R (%) References

Pes

tici

des

Clofibric acid 0.0003 140.8 15.8 not available 81.0 [141]

Fenoprop 0.0013 35.8 49.2 not available 83.0 [141]

Propoxur 0.0026 18.1 73.8 not available 58.0 [141]

Pentachlorophenol 0.0171 2.7 > 99.0 >99.0 [174] 83.0 [141, 175]

Atrazine 0.0078 5.9 98.4 7.0 [168] 4.4 – 36.0 [141, 169]

Ametryn 0.0036 12.9 84.7 not available 92.0 [141]

Indust

rial

chem

ical

s 4-tert-butylphenol 0.0000 - < 0.1 not available 91.0 [141]

Bisphenol A 0.0000 - < 0.1 0 – 32.0 [168, 176] 90.4 [141]

4-tert-octylphenol 0.0000 - < 0.1 not available 96.0 [141]

Phyto

estr

ogen

Formononetin 0.0100 4.6 > 99.0 >99.0 [177] 93.0 [141]

Enterolactone 0.0202 2.3 > 99.0 not available 92.0 [141]

UV

filt

ers

Benzophenone 0.0000 - < 0.1 not available 99.0 [141]

Oxybenzone 0.1797 0.3 > 99.0 >99.0 [178] 98.0 [141, 178,

179]

Octocrylene 0.0000 - < 0.1 not available 88.0 [141]

Removal efficiency (R), biodegradation rate constant (kb), and half-life time (t1/2).

111

In this investigation, negligible removal was observed for compounds, including

metronidazole, ibuprofen, benzophenone, 4-tert-butylpheol, octocrylene, and

bisphenol A. Previous studies examining aerobic conditions, nonetheless, reported

moderate to high removal of these TrOCs [141, 169, 176, 179-181]. The difference

between anaerobic and aerobic microbes can be a possible explanation for this

observation, suggesting higher capacity of degrading TrOCs of aerobic populations.

This could be clarified by identifying microbes involved in anaerobic digestion,

which was however beyond the scope of this study.

6.3.2 Role of chemical structures

The bioavailability of trace organics during anaerobic conversion depends on not

only distinct bacterial community and their synergic effects but also their

physicochemical features. Therefore, an assessment of the relative effects of TrOC

properties on their removal efficiencies was more focused in order to establish

certain applicable generalisations.

Results in this batch test exhibited the varied removal extent by anaerobic

biodegradation in both hydrophilic and hydrophobic compounds. Under aerobic

conditions, it was reported that the excellent removal were observed in hydrophilic

compounds (log DpH = 8< 3.2) [141, 169, 180]. This is because most of those aerobic

studies examined the decrease of TrOC concentrations only in the water phase. As

such, two main mechanisms, namely biodegradation in the water phase and sorption

from the water phase into solid phase were responsible for this change. On the other

hand, the removal of TrOCs in the current investigation was determined from their

concentrations in both the water and solid phases. Thus, biodegradation was the only

removal mechanism. Since these TrOCs were selected based on their diversity with

respect to origins, intended usages, and physicochemical features, it is essential to

examine the role of chemical structures in determining their biodegradability under

anaerobic sludge treatment.

A detailed examination of the chemical structures and properties of the selected

TrOCs was conducted to elucidate their biodegradability under anaerobic condition.

When chemical features in terms of the complexity of aromatic rings and properties

of functional groups were compared, it is noted that there was no clear association

between the nucleus complexity and the elimination capacity of selected TrOCs. This

112

is because the main tendency of bacterial activity is mostly to breakdown any

compound from exterior structure, followed by a further nucleus attack.

Considering electron donating (EDG) and electron withdrawing (EWG) groups play

a discernible role in electrophilic nucleus orientation, a great deal of research has

indicated greater effect of EDG than EWG on making organic compounds more

susceptible to biodegradation. The underlying rationale was that EDG is inclined to

drive molecules more susceptible to electrophilic attack by oxygenases of aerobic

microbes. This apparent correlation was successfully utilised to create feasible

frameworks for predicting the removal extent of given sets of TrOCs during the

aerobic treatment processes [141, 169, 182]. Nonetheless, this framework is not

suitable for assessing the biodegradability of TrOCs under an anaerobic condition.

Table 25 summarises key functional moieties that can influence the removal

efficiency of TrOCs in an anaerobic condition.

Anaerobic digestion showed excellent removal of formononetin, oxybenzone, and

naproxen (>99%). The inclusion of methoxy group in their structures could be

responsible for complete bioconversion of these three compounds. As previously

discussed, the readily biodegradability of methanol solvent was demonstrated to

enhance methanogenesis in TrOC-spiked bottles. The growth of specific bacterial

species selectively working on organic compounds analogous to methanol was

consequently stimulated. Therefore, these compounds were simultaneously

biotransformed with methanol from the start-up of digestion (i.e. t1/2 values of

oxybenzone, naproxen, and formononetin are 0.3, 0.4, and 4.6 days, respectively). In

a good agreement with the literature, Liu et al. [178] suggested that oxybenzone was

more favourably degraded by anaerobic incubation than other redox conditions,

achieving its complete removal after 42 days. In a more recent study of UV filters in

aquifers, the well degradation under aerobic and anaerobic conditions of this

compound were also observed by Liu et al. [179]. In order to enhance the

biodegradation of certain organic compounds in the absence of oxygen, the usage of

methanol as an electron donor in anaerobic metabolic pathways has been widely

investigated in the literature. By means of dehalogenation, a nearly complete

elimination of halogenated compounds in methanol-fed anaerobic digesters has been

recorded (i.e. more than 99% for hexachlorocyclohexane [183] and 2,4,6-

trichlorophenol [184]) compared to that of digesters fed with other electron donors.

113

Table 25. Relation of functional moieties and the removal efficiency of the selected

TrOCs.

Compounds containing chloride groups and/or amine/amide with the effective

removal

Fenoprop

Diclofenac

Triclosan

Pentachlorophenol

Atrazine

Primidone

Ametryn

Carbamazepine

Propoxur

Compounds containing methoxy groups with the high effective removal

Naproxen

Formononetin

Oxybenzone

Compounds containing long alkyl chain with low removal efficiency

Ibuprofen

4-tert-

butylphenol

Clofibric acid

4-tert-octylphenol

Octocrylene

Amitriptyline

In this investigation, a varied range of removal efficiencies was observed with

chlorinated TrOCs in the range of 15.8% to more than 99% (Table 25). The

increasing order of removal was observed with the greater extent of chlorination,

with the excellent removal of pentachlorophenol (more than 99%). This observation

is in a good agreement with the literature [175]. The high biodegradability this class

of compounds under anaerobic condition can be attributed to anaerobic metabolism

through the dehalogenation pathway. By contrast, aerobic microorganisms initialize

the degradation pathway of halogenated compounds via oxidation from other co-

114

existing functional groups [169]. This observation suggests that less halogenated

compounds are more readily biotransformed with oxygen exposure. On the other

hand, the inclusion of alkyl chain in chloride-containing compounds could affect

their conversion pathway, resulting lower removal efficiencies, for which clofibric

acid was a particular example (15.8% elimination).

The presence of amide/amine groups in TrOC chemical structures can affect their

removal under anaerobic condition. Some of TrOCs bearing amine/amide have been

shown to be poorly removed in previous studies. Carbamazepine was, partly

removed with efficiency of 48.3% in this investigation, considered as a typical

example of very persistent xenobiotic compounds regardless of redox conditions [25,

168, 172, 173]. The moderate removal (38.5%) was here recorded for primidone,

which was found to pass through soil to groundwater under anaerobic condition by

Ternes et al. [172]. Higher removal of such compounds as diclofenac and atrazine

(60% and 85%, respectively) may be also due to the co-existence of the chloride

group in their chemical structures, which was poorly removed by aerobic MBR

treatment [141, 169]. Propoxur and ametryn also presented high removal efficiencies

(i.e. 73.8% and 84.7%, respectively) as reported [141, 169], showing their highly

bioavailability to any redox conditions. Interestingly, amitriptyline bearing both

chloride and amine groups showed highly persistent to anaerobic degradation. This

behaviour could be ascribed to the existence of long alkyl group.

Negligible removal efficiency (kb ≈ 0 L/gVS.d) was observed with some compounds,

including ibuprofen, 4-tert butylphenol, 4-tert-octylphenol, octocrylene, bisphenol A,

amitriptyline, and benzophenone. Except for benzophenone, which has no functional

group, the inclusion of alkyl group (Table 25) in chemical structure of other

compounds was responsible for their poor biodegradation. It should be noted that

more than 70% dissipation of octocrylene was obtained after 77 incubation days with

anaerobic consortium in a native aquifer [179]. These contrasting observations here

may be associated to the distinct indigenous bacteria populations available in

aquifers and anaerobic sludge. To our best understanding, while investigations

regarding the biodegradation of 4-alkylphenols under anaerobic condition have been

still limited. Their great attenuation [180] and involved biodegradation pathways

[181] have been only studied in the context of aerobic processes. The inclusion of the

alkyl side chain was probably responsible for their poor removal. In the literature, a

115

negative effect of long alkyl chain on the biodegradation rate was demonstrated in

case of phthalate esters [185]. The concentration of bisphenol A seemed to be stable

during this anaerobic conversion. Bisphenol A was previously reported to be

insignificantly degraded by anaerobic consortium at efficiency of 37% [168] and

even persistent to anaerobic degradation [176]. A significant decrease in

concentration of this compound, by contrast, has been found under different aerobic

treatment processes [176, 180].

For steroid hormones, it is noted that the high biodegradation (>83.5%) was achieved

with 17-β-estradiol, while estrone exhibited the lower removal efficiency (62.0%) at

day 35. This observation could be attributed to possible oxidation of 17-β-estradiol to

estrone, which was observed in anaerobic sediments [23, 186]. The various extent of

biodegradation of members in this group has been widely but contrastingly reported.

A substantial level of estrone, and 17-β-estradiol was still detected in UASB effluent

whilst any type of sludges did not show the removal capacity of 7-α-ethinylestradiol

[126, 186]. By contrast, Carballa et al. [25] reported significant decreases in sum

concentrations of estrone and 17-β-estradiol (85%), and 17-α-ethinylestradiol (60%),

whose removal in this batch test was reported at 45%. Little was known about the

anaerobic biodegradation of 17-β-estradiol-17-acetate, which was here observed to

be effectively degraded (> 99%).

The existence of functional groups in chemical structures of TrOCs determined their

various biodegradation rates throughout the anaerobic conversion. In particular,

while the inclusion the halogen, methoxy and amine/amide groups rapidly simulated

the anaerobic degradation/transformation of TrOCs; the alkyl group was responsible

for their biological recalcitrance to anaerobic treatment.

6.4 Fate of trace organic compounds in sludge matrix under anaerobic

condition

The distribution of TrOCs in the water and solid phases during anaerobic treatment

can be presented by kp obtained from the two-phase fate model. Within anaerobic

treatment of sewage sludge, both kp values reported here and those reported in the

literature showed comparable sorption properties of respective TrOC groups.

Negligible adsorption (kp< 0.1 L/g) of pharmaceuticals such as naproxen, ibuprofen,

and diclofenac to the solid phase reported in the current investigation is consistent

with previous studies in case of primary and secondary sludge [187] and of digested

116

sludge [188]. A correspondence between high kp values and their high

hydrophobicity nature was only found in steroid hormones, of which 17-β-estradiol-

17-acetate was an exception. For the other compounds in this group, there is an

agreement of their high adsorption potential between this study and previous studies,

estimated kp values here were lower than values reported from the literature [189].

In the initial period of spiking, a fraction of investigated compounds was adsorbed

onto solid particles. Different sorption properties onto sludge particles were found

for these trace organics over the experimental time.

It was expected that the distribution of TrOCs was in a relation with their

physicochemical properties, particularly their hydrophobicity (log D). Figure 39,

however, shows a weak correlation between log D (pH = 8) and log kp of the TrOCs

investigated here (R2 = 0.427). The same observation was stated in the literature

although several previous attempts have been made to analyse the relationship

between log D and adsorption of substances into different environmental samples,

including activated sludge [190], primary and secondary sludge [187], and digested

sludge [188]. Even though pKa at the ambient pH was considered in modelling,

Carballa et al. [188] found a significant high deviation modelled values in case of

iopromide, sulfamethoxazole and roxithromycin. One possible explanation is that

sorption behaviour while depends on physicochemical properties of involved

chemicals and solid compositions, is still affected by other conditions, including pH,

temperature, and ion strength [191]. Sorption behaviour was also found to be

dependent on initial concentrations applied as reported in case of bisphenol A [189].

Another reason is that biodegradation factor was taken into account. As such, in view

of diverse TrOC groups selected in this anaerobic treatment study, their

hydrophobicity and distribution in two phases of sludge was then examined along

with different biodegradation rate.

117

-2 0 2 4 6 80.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

2.4

Linear regression

R2

= 0.427

log

kp

log D at pH =8

Figure 39. Correlation between log D (pH = 8) and log kp of the selected TrOCs.

The biodegradability of TrOCs is classified based on biodegradability rate (kb) as

illustrated in Figure 40. For compounds with negligible to low removal rates (kb ≤

0.001 L/gVS.d), their hydrophobicity (log D at pH = 8) and their partition (kp) in

sludge were comparable. There was a moderate relationship (R2 = 0.715) between

them as illustrated in Figure 41. In this occasion, the biodegradation rate had no

effect on their distribution in sludge system. Instead, the adsorption tendency of these

TrOCs was dependent on their hydrophobicity (known as log D). High

concentrations in the solid phase (kp> 0.1 L/g) of some compounds such as

octocrylene and 4-tert-butylphenol could be attributed to the inclusion of the alkyl

group, which referred to biological persistent TrOCs (Section 6.3.2). Similarly,

Ismail et al. [192] observed an increase of sorption capacity according to increasing

alkyl chain length in case of quaternary ammonium compounds. Therefore, the fate

in two-compartment sludge of biological persistent compounds could be predicted by

means of their hydrophobicity without any effect of biodegradability.

118

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log D kp

kb

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ter-s

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arti

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kp

)

log

D (

pH

= 8

)

0.00

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0.16

0.18

0.20

0.00

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0.16

0.18

0.20

Bio

deg

ra

da

tio

n r

ate

(k

b)

kb < 0.001 L/gVS.d kb > 0.001 L/gVS.d

Figure 40. Relation among biodegradation rate constant (kb, L/gVS.d) water-solid

partition (kp, L/g) and log D (pH =8) of the selected TrOCs.

For the other compounds, which are amenable to anaerobic digestion (kb> 0.001

L/gVS.d), their biodegradability rate and sorption properties played an important role

in governing their elimination in anaerobic sludge treatment. There was a weak

correlation between log D (pH = 8) and kp of these TrOCs (Figure 41) due to low

correlation coefficient (R2 = 0.301). Log D values should be ruled out from

explaining their fate in two-compartment sludge under anaerobic condition in this

case.

119

-2 -1 0 1 2 3 4 5 6 7 8

0.00

0.05

0.10

0.15

kp

Linear regression

R2

= 0.715

log D at pH =8

a. kb< 0.001 L/gVS.d

-2 -1 0 1 2 3 4 5 6

0.00

0.05

0.10

0.15

0.20

k

p

log D at pH =8

Linear regression

R2

= 0.301

b. kb> 0.001 L/gVS.d

Figure 41. Correlation of log D (pH = 8) and kp of compounds showing low (a) and

high (b) biodegradability rates (kb).

120

6.5 Summary

System stability of this anaerobic sludge digestion was confirmed by evaluating

temporal profile of control parameters and the CH4 production. Firstly, favourable

operating conditions in case of pH and buffer capacity (i.e. close to neutral pH and

high alkalinity of more than 4000 mgCaCO3/L, respectively) were observed in

TrOC-spiked bottles throughout the digestion time. Secondly, results of CH4

production obtained from experimental data and from the first order and modified

Gompertz kinetic models showed mostly similar CH4 generation between bottles

with and without TrOCs. Their differences, including higher CH4 production from

day4 to 9 and longer lag time of around 1.6 days in spiked bottles, resulted from the

usage of methanol as the solvent in TrOC preparation. A steady state of

methanogenesis was then created in TrOC-spiked bottles after 1.6 days.

Their various removal efficiencies were consistently related to their molecular



exhibited an effective removal while lower to negligible elimination efficiencies

were observed in compounds possessing long alkyl group under same conditions. In

this study, the main removal mechanism of investigated TrOCs from sludge matrix

was anaerobic biodegradation. In terms of TrOC removal, the distribution of selected

TrOCs in water and solid phase was examined using the two-phase fate model.

Regarding compounds possessing high biodegradability (kb> 0.001L/gVS.d), kp

values instead of their hydrophobicity (namely log D) calculated from the proposed

two-phase fate model demonstrated a relation to their partition in the sludge phase.

For biologically persistent compounds to anaerobic conversion (kb< 0.001L/gVS.d),

sorption properties, which can be predicted by log D values, were responsible for

their fate of in sludge matrix.

121

7 CONCLUSIONS AND RECOMMENDATIONS

7.1 Conclusions

The treatment performance of sewage sludge under anaerobic condition was

investigated in this study. Firstly, a critical review on the up-to-date literature of the

AD process was concentrated on two key aspects, namely CH4 production

enhancement and TrOC removal efficiencies. Secondly, a series of batch tests with

respect to both aspects were conducted to envision the potential of glycerol as a co-

substrate in enhancing CH4 yields, and to widen the understanding of the capacity in

eliminating diverse TrOCs in sewage sludge under anaerobic condition. The main

conclusions were summarised as following.

In the first series of batch-mode biomethane potential (BMP) tests, the potential of

CH4 yields and process stability of each sewage sludge and glycerol was tested based

on their characteristics. Data from experimental observations indicated a significant

improvement in terms of yielding CH4and system stability by the buffer supplement

(NaHCO3). Compared to the non-buffed bottles, higher performance in producing

CH4 was recorded in bottles with addition of 15 and 30 mM NaHCO3. Estimated lag

times from kinetic analyses of the first-order and modified Gompertz models also

gave an evidence of such stimulation of CH4 generation by buffer addition. It is

noteworthy that buffer concentrations applied in this study were still not adequate to

initialise CH4 production. This instability could be attributed to insufficient

methanogenic activity.

BMP of raw primary sludge was alternatively tested by using 1/1 ratio rather than 1/9

ratio of I/S by volume. Stability in methanogenesis of raw primary sludge was

achieved during this batch test due to favourable anaerobic conditions (i.e. neutral

pH and high buffer capacity), revealing an important role of I/S ratio. Both

experimental results and kinetic analysis data demonstrated a great potential of raw

primary sludge in CH4 yield, low CH4 production rate and long lag phase made it

unfavourable for a well-established anaerobic digester though. The supplement of

substrates rich in readily biodegradable part was prospectively considered to

accelerate the methanogenesis of raw primary sludge in particular and sewage sludge

in general.

122

Another series of BMP tests of glycerol were carried out with the adequate inoculum

supplementation. Evaluation of control parameters among three types of glycerol at

different concentrations showed a stable function of anaerobic conversion. It could

be expected from the high experimental and predicted CH4 potential yields that any

glycerol should be considered a feasible co-substrate for sewage sludge. Their high

soluble organic fraction and solubility also supported this expectation.

Data from batch tests of anaerobic co-digestion of raw primary sludge and glycerol

pointed out that the addition of 0.5% and 1.0% of glycerol into raw primary sludge

was satisfactory in terms of daily and cumulative CH4 production within the start-up

stage. This improvement resulted from the enhancement of organic source with

regard to CODs, the highly biodegradability and solubility of glycerol. On the other

hand, certain characteristics of digestates, typical of higher CODs, acidic pH and

inadequate alkalinity as well as short methanogenic duration, indicated the

incomplete anaerobic biodegradation for the long-term operation. These findings

indicated the great effects of I/S ratio and buffer supplementation on system stability

and then the ultimate CH4 potential.

The last batch-mode BMP tests were performed to evaluate the removal efficiency of

selected TrOCs by anaerobic treatment. In terms of system stability, experimental

results and kinetic data of parameter variations and CH4 yields indicated a steady

state of methanogenesis, which was created after 1.6 days in TrOC-spiked bottles.

Biodegradation was the main mechanism for the removal of TrOCs during this AD

of sludge. The extent of biodegradability of TrOCs was related to their molecular



exhibited an effective removal while low to no removal efficiencies were observed in

compounds possessing long alkyl group under same conditions. In general, the

distribution of selected TrOCs in sludge system under anaerobic conditions exhibited

a weak correlation to their hydrophobicity (log D). Particularly considering

biologically compounds persistent to anaerobic conversion, their water-solid partition

(kp) can be predicted by their log D. In case of TrOCs amenable to anaerobic

conversion, kp calculated from the proposed two-phase fate model are more

appropriate to describe their fate in sludge matrix than log D values.

123

7.2 Recommendations for future studies

Some suggestions for future studies were made over this research course.

Given the great effects of buffer supplement and I/S ratio on the system stability, it is

highly recommended to investigate appropriate I/S ratios and buffer concentrations

for a stable performance of methanogenic conversion, after which the ultimate CH4

yields of co-substrate mixtures will be attained.

It would be also necessary to evaluate the limit glycerol concentration beyond which

a shock of organic loading will happen under steady state of anaerobic processes.

Further, BMP of anaerobically co-digesting sewage sludge with other potential

organic materials, beverage rejects and food waste, as examples should be

extensively investigated. These findings at laboratory-scale batch mode are expected

to be valuable for larger scale operations in WWTPs in terms of economic benefits

and time saving.

Compounds possessing the methoxy group in their chemical structures effectively

eliminated with the presence of methanol under anaerobic condition, suggesting in

some extent a positive role of methanol in anaerobic biodegradation of TrOCs. This

suggestion deserves more studies to comprehensively examine a real effect of co-

digestion with readily biodegradable substrate on anaerobic metabolism of a

particular compound.

The observed significant variation from negligible to excellent removal capacity of

the anaerobic sludge treatment requires more understanding of main factors, which

govern the elimination of specific chemicals. The long-term exposure of inoculum to

TrOCs has been considered as an example. This aspect, however, has not been

included in this study. Such acclimatisation could be possibly envisioned through

more batch experiments using TrOC-acclimatised anaerobic sludge. Quantification

and identification of associated anaerobic populations responsible for degrading

TrOCs could be also endorsed in a broader context.

Modifications of the applied two-phase fate kinetic model should be made to develop

a novel mathematic model. Considering all possible factors associated to the

dynamics of TrOC concentrations over the digestion time, this model is expected to

more precisely predict their fate in sludge matrix under anaerobic condition, and their

attenuation subsequently.

124

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135

APPENDIX

Table 26.Selected TrOCs and their relevant physicochemical properties.

Group Compound Molecular

formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

Phar

mac

euti

cals

Salicylic acid C7H6O3 138.12 –1.14 1.42 × 10-8

Ketoprofen C16H14O3 254.30 –0.55 1.92 × 10-13

Naproxen C14H14O3 230.26 –0.18 2.07 × 10-12

Metronidazole C6H9N3O3 171.15 –0.14 6.08 × 10-12

136


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

Ibuprofen C13H18O2 206.28 0.14 5.54 × 10-12

Primidone C12H14N2O2 218.25 0.83 1.16 × 10-14

Diclofenac C14H11Cl2NO2 296.15 1.06 2.69 × 10-11

Gemfibrozil C15H22O3 250.33 1.18 1.83 × 10-11

Carbamazepine C15H12N2O 236.27 1.89 9.41 × 10-12

137


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

Amitriptyline C20H23N 277.40 3.21 1.24 × 10-10

Triclosan C12H7Cl3O2 289.54 4.92 9.49 × 10-6

Ste

roid

horm

ones

Estriol C18H24O3 288.38 2.53 1.75 × 10-11

Estrone C18H22O2 270.37 3.62 9.61 × 10-10

138


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

17-α -ethinylestradiol C20H24O2 269.4 4.11 3.74 × 10-10

17-β-estradiol C18H24O2 272.28 4.15 1.17 × 10-9

17-β-estradiol-17-acetate C20H26O3 314.42 5.11 2.15 × 10-9

Pes

tici

des

Clofibric acid C10H11ClO3 214.65 –1.29 2.91 × 10-10

Fenoprop C9H7Cl3O3 269.51 –0.28 4.72 × 10-12

139


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

Propoxur C11H15NO3 209.24 1.54 5.26 × 10-7

Pentachlorophenol C6HCl5O 266.34 2.19 1.82 × 10-7

Atrazine C8H14ClN5 215.68 2.64 5.22 × 10-8

Ametryn C9H17N5S 227.33 2.97 3.67 × 10-9

140


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

Indust

rial

chem

ical

s

4-tert-butylphenol C10H14O 150.22 3.39 7.51 × 10-6

Bisphenol A C15H16O2 228.29 3.64 1.34× 10-9

4-tert-octylphenol C14H22O 206.32 5.18 8.67 × 10-6

Phyto

estr

ogen

Formononetin C16H12O4 268.26 1.81 2.91 × 10-10

Enterolactone C18H18O4 298.33 1.88 8.07 × 10-13

141


formula

Molecular weight

(g/mol)

Log D at

pH = 8

Henry’s Law

constant at

25 °C

(atm.m3/mol)

Chemical structures

UV

fil

ters

Benzophenone C13H10O 182.22 3.21 1.31 × 10-6

Oxybenzone C14H12O3 228.24 3.42 1.22 × 10-8

Octocrylene C24H27N 361.48 6.89 3.38 × 10-9

All values were sourced from Scifinder Scholar database (ACD/Labs)

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